BIOLOGY AND CONSERVATION OF MARINE TURTLE NESTING IN … · B Dew, Dr. MA Ditmer, Dr. M Godfrey, Dr. S Kross, Dr. N Mrosovsky and Dr. S Weber. Quisiera agradecer a mi profesora de

BIOLOGY AND CONSERVATION OF MARINE TURTLE NESTING IN THE DOMINICAN REPUBLIC

Tesis Doctoral por: Ohiana Revuelta AvinDirectores: Jesús Tomás Aguirre / Juan Antonio Raga Esteve

Valencia, mayo 2014

PROGRAMA DE DOCTORADO EN BIODIVERSIDAD 3001

BIOLOGY AND CONSERVATION OF MARINE TURTLE NESTING IN THE DOMINICAN REPUBLIC

Tesis Doctoral por: Ohiana Revuelta AvinDirectores: Jesús Tomás Aguirre / Juan Antonio Raga Esteve

Valencia, mayo 2014


D. Jesús Tomás Aguirre, Profesor Ayudante Doctor del departamento de Zoología de la Facultad de Ciencias Biológicas de la Universitat de València y D. JuAn AnTonio rAgA esTeve, Catedrático del departamento de Zoología de la Facultad de Ciencias Biológicas de la Universitat de València,

CERTIFICAN que Dª Ohiana Revuelta Avin ha realizado bajo nuestra dirección, y con el mayor aprovechamiento, el trabajo de investigación recogido en esta memoria, y que lleva por título “BIOLOGY AND CONSERVATION OF MARINE TURTLE NESTING IN THE DOMINICAN REPUBLIC”, para optar al grado de Doctora en Ciencias Biológicas.

Y para que así conste, en cumplimiento de la legislación vigente, expedimos el presente certificado en Paterna, a de de 2014.

Firmado: Jesús Tomás Aguirre Firmado: Juan Antonio Raga Esteve

A Héctor

AGRADECIMIENTOS En primer lugar quisiera agradecer a mis directores de tesis por darme la oportunidad en su día de seguir trabajando con tortugas y poder realizar esta tesis doctoral. A Jesús, que confió en mí para coordinar el nada fácil trabajo de campo en República Dominicana. Gracias por confiar entonces, por todo el apoyo y por creer en mí para hacer esta tesis, ayudándome tanto durante el largo proceso. Gracias Toni por ayudarme con todo, por estar pendiente y ponerme las cosas fáciles tanto en República Dominicana como en Valencia.

Quiero mostrar mi agradecimiento y admiración a Yolanda León, de quien he aprendido no solo a nivel científico (que ha sido mucho) sino también a nivel personal. Gracias por recibirme en tu casa y hacerme un hueco en tu vida, por ayudarme en todo momento y enseñarme tantas cosas. Tengo que agradecer a toda la familia León por acogerme en su casa y hacerme sentir como una más. A todas las amigas de “la doctora” que tuvieron que soportarme en sus cenas y reuniones… con especial cariño a Ana Laura y Gabi, con quienes compartí más tiempo. Gracias también a Denise y a Santiago por su amistad.

Tengo tanto y tanta gente a quien agradecer en el sur profundo… Quiero dar las gracias sobre todo a mis compañeros de trabajo de campo en las playas del Parque Nacional Ja-ragua, por el trabajo duro que hicieron, por compartir caminatas, mosquitos, “cadillos”, noches en vela y demás penurias, muchas veces sin entender muy bien para qué. Gracias sobre todo a “Pablito” (Pablo Feliz) y a “Blanco” (Bienvenido Pérez Turbí), compañeros in-fatigables, por su amistad y por su apoyo y, como no, gracias también a Macuso, Francisco y Quico, siempre dispuestos a acompañarme. No puedo olvidarme de Andy, compañera de casa, muestreos y mil aventuras, me alegro de haber compartido ese tiempo contigo. I also have to thank Kate and Stefania for their invaluable help in the fieldwork in 2009. Gracias a tantos otros y otras que en algún momento pasaron por Bahía de las Águilas y pusieron su granito de arena en la conservación de las tortugas marinas colaborando en los recorridos de playa. También quiero dar las gracias a la familia de “Doña Chava” con quien viví y compartí mi primera estancia en República Dominicana. Gracias a toda

la gente que pasó por Chava en esa época, con todos ellos compartí interesantes veladas, charlas y excursiones a lugares increíbles. A la comunidad española de Pedernales con quien pasé buenos ratos, sobre todo, gracias a Jose por las charlas, las caminatas y las “presidente” en la plaza y el malecón.

Todo mi cariño y gratitud para el equipo de trabajo “Saonero”, pasamos tantas noches playa arriba, playa abajo buscando “careyas”… Sobre todo tengo mucho que agrade-cerle a “Negro” (Pelagio Paulino), no solo por ayudarme en el trabajo (o más bien por dejar que yo le ayudara a él), sino por adoptarme y cuidar de mi durante mis estancias en la isla. Gracias por tu ánimo, por ser inagotable, por tener siempre mil ideas y por tu cariño. Atesoro mil historias y vivencias de mi estancia en Saona y de su gente, no puedo nombrar a cada uno de sus habitantes, pero siempre recuerdo con mucho cariño a toda la comunidad de Mano Juan, especialmente a las “doñas” y a los niños con quienes pasé tantos buenos ratos. Tengo que dar especialmente las gracias a Miledy por su hospitalidad y a “Mico” por su cariño y su buen humor y, como no, a todo el equipo tortuguero: Yoval, Dariel, “Chino” y Vinicio por su duro trabajo.

Gracias a toda la gente del grupo Jaragua por todo el apoyo que recibí siempre, especial-mente a Yvonne Arias, por su ayuda, y a Dawaira y Laura por su alegría y amistad.

Gracias a toda la gente de la Unidad de Zoología Marina con quienes he compartido principalmente en los dos últimos años. Especialmente quiero agradecerle a Javi todo lo que me ha ayudado a lo largo de esta tesis, gracias por tener siempre la puerta de tu despacho abierta y una sonrisa, ya fuese para resolver una duda estadística, gramatical, ornitológica o ¡existencial! Gracias por tu apoyo, por confiar en mí y animarme siempre a seguir. A Juan Antonio por resolverme muchas dudas, por animarme a conocer el “mun-do R” y compartir sus conocimientos conmigo; gracias también por tu confianza y por tu sentido del humor. A Merche con quien he disfrutado comidas, risas y divertidas salidas pajareras, gracias por todo. Tengo mucho que agradecerle a Patri, por encomendarme y enseñarme el trabajo del barco (mis primeros intentos con GIS); gracias por tenerme en cuenta para muchos otros trabajos y por ser tan buena gente y estar siempre dispuesta a ayudar. Gracias a Natalia por ser tan “Lady” y tan “beia”, porque nunca está demasiado

ocupada para echar una mano, gracias por estar siempre ahí para lo que haya necesitado. A la “tortuguera uruguaya”, por la sonrisa perenne, los ánimos, la ayuda y el buen rollo, muchas gracias Gabi. A Raúl, compañero de máster, rutas de domingueros, pajareo, cur-sos de fotografía y lo que nos echen… gracias por tus dosis de tranquilidad, tu ayuda y tu amistad. Ana Born, compañera de penas y risas doctorales, gracias por todo y mucho án-imo en la recta final. Al otro “tortuguero-montañero”, gracias Francesc por ser tan buen compañero y por estar siempre dispuesto a echar una mano. Al resto de compañeros con quienes he compartido comidas, cafés, dudas, risas, muchas gracias a todos por todo: al dúo Neus-Ana Ahuir por sus risas, a Abril por todas sus locuras, a Paula, Aigües, Eugenia, Isa Alonso y Ana Pérez por las variopintas “charretas” compartidas. A Paco por tener siempre una frase alegre de buena mañana. Gracias también a los que estuvieron, a las nuevas generaciones y otros compañeros eventuales con los que también compartí mo-mentos de despeje y risas alrededor de la máquina de café: gracias Celia, Cristina, Gema, Isa Blasco, Jesús H., Jose, Mar, María, Salvatore, Wolf. A mis compañeros de máster por todas las horas invertidas en tratar de arreglar el mundo, hacer trabajos, y por los ratos de fiesta, muchas gracias sobre todo a mis compis del “ala celta”.

En la Universidad de Valencia tengo que agradecer también al profesor Pepe Pertusa por hacerme un hueco en su laboratorio, por su ayuda y consejos. A Paco Ruiz, gracias por dedicarme tu tiempo cuando todo parecía imposible con el ArcGIS. También tengo que agradecerle mucho a Pascual Asensi, gracias por salvarme más de una vez con la in-formática, por tener siempre tiempo para echar una mano.

From the Exeter University I specially thank Dr. Brendan Godley, who I consider the third supervisor of this thesis. I’m in debt for all your help, thank you for your patience, guidance and always constructive comments. I would also like to thank Dr. Lucy Hawkes, who patiently helped and solved my doubts working with satellite data, and Dr. Annette Broderick for her valuable collaboration in several parts of the study.

Tengo que agradecer a todos los investigadores, tanto de tortugas marinas como de otras disciplinas, que dedicaron parte de su tiempo a responder mis correos cargados de dudas y que, de manera totalmente altruista, colaboraron con ideas y soluciones. Thanks to Dr.

B Dew, Dr. MA Ditmer, Dr. M Godfrey, Dr. S Kross, Dr. N Mrosovsky and Dr. S Weber.

Quisiera agradecer a mi profesora de inglés Candice su ayuda y dedicación a la revisión del inglés de buena parte de esta tesis, thank you Candice for your generous help, I’m really grateful. A los fotógrafos dominicanos Ricardo Briones (Ricky) y José Alejandro Álvarez por responder de inmediato y con tanto cariño a mi petición de material gráfico, poniendo a mi disposición muchas de las increíbles fotos de esta tesis.

Por supuesto quiero agradecer el apoyo, los ánimos, los empujones en los momentos bajos, las risas, las lágrimas, las cervecitas, las reuniones, las salidas al monte, las charlas y todo lo que hemos compartido durante esta tesis a mi familia y amigos. Es imposible (bueno, más bien un rollo…, lo reconozco) nombrar aquí a todos, pero esto es para todos vosotros que habéis estado ahí. Ha sido un camino muy largo y en las diferentes etapas ha habido gente más cerca y otros más lejos, pero siempre me he sentido muy apoyada. Mil gracias por eso, todo hubiese sido mucho más difícil si no hubieseis estado ahí. Quiero agradecer especialmente a mi familia valenciana, gracias a ellos mi llegada a Valencia fue mucho más fácil, gracias a todos por recibirme con los brazos abiertos, por alterar vuestras vidas y dedicarme tanto tiempo; sobre todo tengo que agradecerle a Juana Mari y a Ángel por hacerme un hueco en casa y hacerme sentir como si fuera la mía, muchísimas gracias.

Una de las personas que más se ha volcado conmigo en esta tesis ha sido Judit (la judi, judiaaa, sacajugo!), gracias por involucrarte tanto y responder siempre con un sí a mis múlti-ples solicitudes de ayuda a lo largo de estos años, esta tesis es muy tuya también. Pero, ante todo, gracias por haberme apoyado siempre con tanta ilusión y haber sido una constante fuente de ánimos y positivismo, tengo mucha suerte de tenerte como amiga.

A mis padres, no sé por dónde empezar a darles las gracias porque necesitaría más de lo que ocupa esta tesis para escribir todo lo que tengo que agradecerles. Nada de lo que he hecho hubiera sido posible sin sus consejos, apoyo y sobre todo sin su cariño, su confianza y respeto. Gracias por enseñarme a ser valiente y a luchar cada día por aquello en lo que creo. Como no, a mi hermana… por todo, por ser la mejor.

Debería otorgarse un título a las personas que, en la sombra, hacen el trabajo sucio de las tesis doctorales… Esta tesis existe porque cada vez que me agobiaba, o el ordenador no funcionaba, o los modelos no salían, o los “referees” me amargaban el día o… tan-tas y tantas quejas y lágrimas (¡ni con R podría calcularlas!), Héctor siempre estuvo ahí, dándome ánimos y apoyándome en todo. Aquí está, lo conseguimos, la terminamos, en-horabuena y gracias, sin ti, ni esta tesis ni nada sería posible.

Y lo prometido es deuda, siempre he dicho que tengo mucho que agradecer a los autores de la banda sonora de esta tesis… gracias a Radiohead por todas las veces que su música me ha acompañado en las diferentes fases de este trabajo.

MARCO Y AGRADECIMIENTOS INSTITUCIONALESEsta tesis doctoral se enmarca dentro del estudio a largo plazo sobre la conservación de las tortugas marinas nidificantes en la República Dominicana, llevado a cabo entre 2006 y 2011 y coordinado por la Universitat de València. Este estudio ha sido posible gracias a la financiación de los siguientes proyectos concedidos a miembros dal equipo de la Uni-dad de Zoología Marina del Instituto Cavanilles de biodiversidad y biología evolutiva: (1) “Estudio de las poblaciones de tortugas marinas nidificantes en el Parque Nacional Jaragua (República Dominicana)” (A/2991/05), según resolución de la Agencia Española de Cooperación Internacional (AECI). Este proyecto fue renovado en 2007, mediante la concesión al mismo equipo del proyecto: A/5641/06 - “Estudio de las poblaciones de tortugas marinas nidificantes en el Parque Nacional Jaragua (República Dominicana) II”. (2) Proyecto nº CGL2006-02936-BOS (Plan Nacional I+D+I, Ministerio de Educación y Ciencia), prorrogado hasta septiembre de 2010. (3) Proyecto GVACOMP2009-331 “Es-tudio de las poblaciones de tortugas marinas nidificantes en el Parque Nacional de Jaragua (República Dominicana)”. 2009. Conselleria de Educación, Generalitat Valenciana. (4) “Conservación de las tortugas marinas en la República Dominicana: relación con el de-sarrollo sostenible de las comunidades locales”. I Convocatòria de projectes de cooperació al desenvolupament de la Universitat de València, Fundación General de la Universitat de València, 2007; proyecto renovado en la II Convocatòria de projectes de cooperació al desenvolupament de la Universitat de València, de la Fundación General de la Universi-tat de València en 2009. (5) Postdoctoral Marie Curie Intra European Fellowship (Union Europea), concedida al Dr. Jesús Tomás para el periodo 2007-2009, desarrollada en la Universidad de Exeter, Reino Unido (Propuesta: “Integrated Ecology of sea turtles”). (6) Marie Curie - Reintegration-Grant FP7-PEOPLE-2009-RG (Union Europea), concedida al Dr. Jesús Tomás para el periodo 2010-2013, desarrollada en la Universitat de València (Propuesta: “Long Term Research on Sea Turtle Ecology and Conservation”).

Estos proyectos han supuesto un esfuerzo coordinado y conjunto entre la Universidad de Valencia, la ONG dominicana Grupo Jaragua, la Universidad de Exeter (Reino Unido), la Universidad INTEC de Santo Domingo, la Universidad Autónoma de Santo Domin-

go (a través del Centro de Investigaciones de Biología Marina - CIBIMA y la Escuela de Biología) y el Programa Araucaria XXI de la Agencia Española de Cooperación Interna-cional en República Dominicana. Quiero agradecer el apoyo y colaboración a todas las personas que han participado en estos proyectos.

Es de agradecer también la colaboración de las siguientes instituciones y personas en estos proyectos: A quienes fueran directores del Programa Araucaria XXI de la AECI en República Dominicana, Juan Enrique García y Rosario Boned y a Marino José y Antonio Trinidad por su apoyo logistico e institucional en Pedernales. A quien fuera director del CIBIMA, Dr. Francisco Geraldes, por su apoyo al principio de este estudio. A la subsecre-taría de Áreas Protegidas, y en especial a José Manuel Mateo, Director de Biodiversidad del Ministerio de Medio Ambiente y Recursos Naturales de la República Dominicana, por su gran apoyo institucional. A The Nature Conservancy por su apoyo en diferentes fases del estudio, especialmente en los trabajos realizados en Saona.

TABLE OF CONTENTS

SCIENTIFIC PUBLICATIONS FROM THE PRESENT PhD THESIS 19

SUMMARY 20

RESUMEN 26

1. CHAPTER I: GENERAL INTRODUCTION 441. 1. Biodiversity loss and conservation 471. 2. What is conservation? 471. 3. Marine biodiversity conservation 481. 4. Coastal and marine protected areas 491. 5. Marine turtles 501. 6. Marine turtles in the Caribbean region 541. 7. Marine turtle conservation: the case of the Dominican Republic 561. 8. The project 63

2. CHAPTER II: AIM AND OBjECTIVES 64

3. CHAPTER III: GENERAL MATERIALS AND METHODS 683. 1. Study area 713. 2. Beach surveys 733. 3. Data collection and clutch relocation 763. 4. Hatching success and emergence success 793. 5. Estimating hatchling sex ratio 823. 6. Tracking hawksbill turtles 823. 7. Statistical analyses 84

4. CHAPTER IV: PROTECTED AREAS HOST IMPORTANT REMNANTS OF MARINE TURTLE NESTING STOCkS IN THE DOMINICAN REPUBLIC 86

5. CHAPTER V: ASSESSING THE EFFICACY OF DIRECT CONSERVATION INTERVENTIONS: CLUTCH PROTECTION OF THE LEATHERBACk MARINE TURTLE IN THE DOMINICAN REPUBLIC 108

6. CHAPTER VI: RUNNING AGAINST TIME: CONSERVATION OF THE REMAINING HAwkSBILL TURTLE (ERETMOCHELYS IMBRICATA ) NESTING POPULATION IN THE DOMINICAN REPUBLIC 128

7. CHAPTER VII: THE VALUE OF ENDANGERED SPECIES IN PROTECTED AREAS AT RISk: THE CASE OF THE LEATHERBACk TURTLE IN THE DOMINICAN REPUBLIC 150

8. CHAPTER VIII: EVALUATING THE IMPORTANCE OF MARINE PROTECTED AREAS FOR THE CONSERVATION OF HAwkSBILL TURTLES (ERETMOCHELYS IMBRICATA ) NESTING IN THE DOMINICAN REPUBLIC 168

CONCLUSIONS 190

REFERENCES 196

Scientific publications from the present PhD thesis I 19

SCIENTIFIC PUBLICATIONS FROM THE PRESENT PHD THESIS

SCIENTIFIC PUBLICATIONS FROM THE PRESENT PhD THESISThe results of this thesis have been pub-lished, accepted or sending for its publica-tion in the following scientific journals:

- Ohiana Revuelta, Yolanda M León, Pablo Feliz, Brendan J Godley, Juan A Raga and Jesús Tomás (2012) Protected areas host important remnants of marine turtle nesting stocks in the Dominican Republic. Oryx, 46, 348-358.

- Ohiana Revuelta, Yolanda M León, Francisco J Aznar, Juan A Raga and Jesús Tomás (2013) Running against time: conservation of the remaining hawksbill turtle (Eretmochelys imbricata) nesting population in the Dominican Republic. Journal of the Marine Bi-ological Association of the United Kingdom, 93, 1133-1140.

- Ohiana Revuelta, Yolanda M León, Annette C Broderick, Pablo Feliz, Bren-dan J Godley, Juan A Balbuena, Andrea Mason, Kate Poulton, Stefania Savoré, Juan A Raga and Jesús Tomás. Assessing the efficacy of direct conservation inter-ventions: Clutch protection of the leath-erback marine turtle in the Dominican Republic. Oryx (In press).

- Ohiana Revuelta, Brendan J Godley, Yolanda M León, Annette C Broder-ick, Pablo Feliz, Juan A Balbuena, Juan A Raga and Jesús Tomás. The value of endangered species in protected areas at risk: the case of the leatherback turtle in the Dominican Republic. Biodiversi-ty and Conservation, doi:10.1007/s10531-014-0682-x.).

- Ohiana Revuelta, Lucy Hawkes, Yolanda M León, Brendan J Godley, Juan A Raga and Jesús Tomás. Evaluat-ing the importance of Marine Protected Areas for the conservation of hawksbill turtles (Eretmochelys imbricata) nesting in the Dominican Republic. Endangered Species Research (Submitted).

OTHER PUBLICATIONS RELATED TO THE PRESENT PhD THESISDuring the course of these studies I have also participated in the following publica-tions within the frame of the marine turtle conservation project in the Dominican Re-public:

- Jesús Tomás, Brendan J Godley, Yolanda M León, Ohiana Revuelta and Juan A Raga (2011) Sea turtle conservation in the Dominican Republic: The impor-tance of National Parks. Testudo, 7, 32-38.

- Lucy A Hawkes, Jesús Tomás, Ohiana Revuelta, Yolanda M León, Janice M Blumenthal, Annette C Broderick, Mar-tin Fish, Juan A Raga, Mathew J Witt and Brendan J Godley (2012) Migratory patterns in hawksbill turtles described by satellite tracking. Marine Ecology Progress Series, 461, 223-232.

- Carlos Carreras, Brendan J Godley, Yolanda M León, Lucy A Hawkes, Ohi-ana Revuelta, Juan A Raga and Jesús Tomás (2013) Contextualising the last survivors: population structure of marine turtles in the Dominican Republic. PLoS ONE, e66037, doi:10.1371/journal.pone.0066037.

SUMMARY©

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22 I BIOLOGY AND CONSERVATION OF MARINE TURTLE NESTING IN THE DOMINICAN REPUBLIC

SUMMARY The loss of biodiversity has become one of the most pressing issues, which has lead to a growing global concern about the status of the biological resources on which hu-man life depends. In recent decades there has been a decrease in individual popula-tions of many species. In the Caribbean, marine turtle’s nesting rookeries have been reduced considerably, mainly due to hu-man exploitation. A number of rookeries in this region have been studied for several de-cades and their status is well documented; however, many other Caribbean rookeries remain poorly described. The Dominican Republic (DR) is an area where informa-tion on marine turtle nesting activity is scarce and outdated. Surveys in the 1970s and 1980s constituted the main reference on the status of marine turtles in the DR; although more recent reports indicate that these species are under continuous threats. The lack of comprehensive studies and re-cent information, coupled with the evident threatened status of marine turtles in the country demanded an updated assessment to help target effective conservation actions.

The present PhD study aims to identify the main marine turtles nesting rookeries in DR, describing the current spatio-tem-poral patterns of nesting, and assessing the likely impact of the current threats to these nesting stocks, based on a period of 5 years (2006 - 2010) of systematic survey within a conservation project. To determine the reproductive success of Dermochelys coriacea and Eretmochelys imbricata in their remnant nesting rookeries in the country: the beach-es of The Jaragua National Park (JNP, SW DR) and Saona Island (in Del Este National

Park: DENP, SE DR) respectively. To face egg take by local people, the project carried out artificial incubation of a number of clutches as a conservation measure at both sites. Differences in hatching and emer-gence success, and sex ratio between arti-ficially and in situ incubated clutches were determined to evaluate the artificial incu-bation programme. High risk of habitat loss was detected on the western beaches of the JNP; hence, the reproductive value of the area for leatherback turtle was determined to reinforce conservation at site. Moreover, the factors affecting hatching success were studied to ascertain what sites can be best candidates to host relocated clutches in case of habitat loss due to development plans. Another aspect of this study focusses on the investigation of the habitat-use patterns of hawksbill turtles during their internesting and foraging periods, identifying core-use areas and comparing them with current marine protected areas in the Caribbean.

The study led to the following findings and conclusions:

For the first time in recent decades we have documented the presence of three marine turtle species (leatherback, hawksbill and green turtle [Chelonia mydas] still nesting in the DR; although the third species was found nesting in very low numbers. Nest-ing is concentrated in the protected areas of JNP and Saona Island, with nesting outside these areas being relatively scarce. The JNP consistently had the highest num-ber of clutches per year of leatherback turtles (mean 126.4 ± SD 74.1, range 17-210), with a total of 632 clutches record-ed during 2006-2010. However, sporadic nesting of hawksbill and green turtles were also recorded. Saona Island hosted the ma-

Summary I 23

jority of recorded hawksbill clutches (mean 100 ± SD 8.4 per year, range 93-111, n = 400) and green turtle clutches (mean 9.2 ± SD 6.2, range 1-15, n = 37). Sporad-ic nesting of leatherbacks and hawksbills was also recorded in 14 other sites in the country. According to the study of clutches, mean clutch sizes (yolked eggs) for leather-back turtle in JNP varied from 67.5 to 75.1 eggs (n = 315 clutches). Mean clutch sizes of hawksbill in Saona Island varied from 125.7 to 139.5 eggs (n = 179 clutches). The leatherback turtle nesting season extends from March to August, with a peak in May, which is consistent with reports for neigh-bouring nesting populations. The season-ality of hawksbill turtle nesting is similar to nearby rookeries on Antigua, Barbados and Mona Island, Puerto Rico, with a high nesting season extending from June to No-vember. Sizes recorded from hawksbills and green turtles nesting in DR were within the normal range of sizes for both species according to literature. The mean size of leatherback turtles recorded in DR (Curved Carapace Length = 147.4 cm) is lower than the global mean, which is normally distrib-uted around 155 cm. This could be the sign of a recovering population with a high pro-portion of neophyte females.

Comparison of these results with earlier re-ports indicates that a profound decline ap-pears to have taken place in the last thirty years. Illegal take of eggs was identified as the main threat to marine turtles, particular-ly on the eastern beaches of JNP. The pres-ent study highlights the need for adequate protection and management of these areas for marine turtle conservation in the DR, since the country seemed to be an important marine turtle nesting area in the past.

In order to face threats, particularly egg take, an official program of artificial incubation was established to protect turtle clutches at JNP and Saona Island. In the JNP, clutches of the western and eastern beaches of the park (called WB and EB respectively), were incubated in hermetic boxes stored in Park rangers’ huts located in both places. The present study included the assessment of effi-ciency of this conservation measure through the investigation of how artificial incubation might be influencing the hatching success of clutches and the resultant sex ratios.

A total of 109 leatherback clutches laid over the study period (2008-2009) were studied: 35 incubated artificially in the EB), 31 artificially incubated clutches in the WB) and 43 clutches incubated in situ at western beaches. The results revealed that the in-cubation method significantly influenced hatching success of clutches. In situ clutches had greater hatching success than those ar-tificially incubated in WB and in EB. The incubation method also had an effect on the number of early stage dead embryos. The number of late stage dead embryos in a clutch was also significantly affected by the incubation method. Artificially incubat-ed clutches in EB and WB had more late stage dead embryos than in in situ clutches. On the other hand, the incubation method significantly influenced incubation dura-tion. Clutches artificially incubated in EB had longer incubation duration than in situ clutches and clutches artificially incubat-ed in WB. Mean incubation temperatures were significantly lower in artificially in-cubated clutches at EB when compared to artificially incubated clutches at WB. Esti-mated sex ratios from artificially incubated clutches at EB (2.5 ± 3.8 % females in 2008


in emergence success between years or in-cubation type compared to in situ incuba-tion. These results indicate that handling was correctly carried out and that incuba-tion conditions in boxes were suitable for hawksbill embryos development. On the other hand, in 21 of the 22 artificially in-cubated clutches mean temperature during the thermosensitive period was lower than the pivotal temperature derived from lab-oratory studies, suggesting a male bias in artificially incubated clutches. Current lev-el of egg take is unsustainable for the long term preservation of this nesting popula-tion. The low percentages of female hatch-ling production inferred from clutch tem-perature data calls for urgent changes in the conservation strategy adopted in Saona.

Studies of in situ hatching success leaded to determine the reproductive value of the western beaches of JNP for the leatherback turtle as well as allowing obtaining data on beach parameters potentially affecting clutch incubation. A total of 64 leather-back clutches spanning three nesting seasons (2007, 2008 and 2009) were studied. The hatching success on these beaches appeared to be driven mainly by the effects of beach sector (i.e, La Cueva beach vs. Bahía de las Águilas beach), the incubation duration, and the date of lay. Hatching success of clutches incubated in the beach sector of La Cueva was higher than in those incubated in Bahía de las Águilas. In addition to beach sector, hatching success was strongly influenced by incubation duration; longer incubation du-rations resulted in lower hatching success. Date of lay affected hatching success, with clutches laid earlier in the nesting season hav-ing higher hatching success. The results of this study show that clutches of leatherback

and 23.4 ± 28.3 % females in 2009) were lower than from those artificially incubated clutches at WB (41.7 ± 23.5 % females in 2008, and 57.7 ± 26.6 % females) and from those in situ clutches (53.6 ± 28.5 % in 2008, and 72.9 ± 30.7 % females). However, since no clutches were incubated in situ in EB we could not ascertain whether there were sig-nificant differences or not on likely sex ra-tios between artificially incubated versus in situ incubated clutches. Clutch relocation is currently the only viable conservation op-tion for clutches on EB due to intense egg take, but steps are needed to ensure that natural sex ratio is not distorted. However, on the WB of JNP, in situ clutch incubation seems possible through beach protection.

Hawksbill nesting population of Saona Is-land also face egg take as a major threat. Hence, a similar official conservation pro-gram, including artificial incubation, was also carried out for hawksbill clutches at Saona. In a four-year period study, a to-tal of 400 hawksbill nests were recorded in 5 sampling areas of nesting beaches at the south coast of Saona. The majority of clutches studied were artificially incubated (n = 163), while 146 clutches had already been predated by humans when found, and the rest were camouflaged and left for in-cubation in situ. However, due to different factors such as egg take or erosion caused by storms, only a total of 49 clutches incu-bated in situ were studied. In total, 12,340 hatchlings were produced under artificial incubations and released to the sea (1731 in 2007, 4522 in 2008, 2664 in 2009 and 3423 in 2010). No significant effects of ar-tificial incubation in hatching success were detected between years or incubation type. Likewise, there was no significant difference

Summary I 25

sion both across turtles and years. The re-sults obtained support previously described internesting behaviour for the hawksbill turtle in other internesting areas in the Caribbean. Overall, the common core-use area during internesting was situated inside the DENP’s maritime boundaries.

Hawksbill turtles tagged in Saona showed a within-rookery dichotomy in their migra-tory strategies, with some turtles remaining close to nesting sites in waters of the DR and others migrating to international foraging grounds. Most of the turtles that migrated internationally foraged in waters off Nicara-gua and Honduras (n = 5). With the excep-tion of two turtles with 2.2% and 91.5% of the tracked days inside a protected area (Sea-flower Biosphere Reserve and Miskitos Cays respectively), the rest of the turtles were in non-protected waters during their entire for-aging period. The two turtles that remained in waters of the DR after nesting in Saona stayed within coastal reef ecosystems. One of these turtles stayed inside the JNP while the foraging area of the second one was located in waters adjacent to Bahía de las Calderas (south DR coast) outside any ma-rine protected area. The remaining tracked turtle travelled northwest to the Bahamas where its core-use area was also out of any protected area. This information highlights the relevance of DR protected areas for the conservation of hawksbill‘s internesting and foraging habitats, showing the need of en-forcing existing legislation for the protected areas in the country. The present study also corroborates that the waters off Nicaragua and Honduras are exceptionally important foraging areas for the species in the Caribbe-an, as well as showing the turtles’ vulnerabil-ity in these waters.

turtles in the JNP western beaches presented unusually high hatching success (75.2%) for this species, compared to other rookeries in the Caribbean and elsewhere. This study is particularly relevant in relation to La Cue-va beach; this sector hosts 20% of the total clutches laid at western beaches of the Park and demonstrated the highest hatching suc-cess levels. However, it is less protected be-cause it is located in the buffer zone out of the Park limits and deserves more protection. Given the exceptional value of hatching suc-cess and the current and potential threats affecting leatherback nesting beaches, addi-tional efforts in regulation and management of the protected area are needed.

Understanding spatial and temporal habi-tat-use patterns to protect both foraging and nesting grounds of reproductive individuals is crucial to success in marine turtle conser-vation. During the conservation project of marine turtles nesting in DR a total of 10 hawksbill females were satellite tagged and tracked after nesting. For the study of the habitat-use patterns we obtained locations from a total of 9 turtles, 8 tagged in Saona Island in August-September 2008 and Sep-tember-December 2009, and 1 in JNP in September 2009. Minimum Convex poly-gon (MCP) analysis showed that individual internesting areas occupied by the turtles ranged from 51.0 to 644.0 km2 (mean ± SD: 254.3 ± 173.6). Nesting hawksbills of Sao-na Island remained in the adjacent waters to their nesting beaches using small home range areas during internesting intervals. Core activity areas occurred in shallow wa-ters mainly within 200 m isobaths and asso-ciated with coral reefs at the eastern-most tip of the island. The home ranges over-lapped, showing similar location and exten-

RESUMEN©

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Resumen I 29

INTRODUCCIóNEs un hecho irrefutable que las entidades que integran la biosfera se encuentran bajo una gran amenaza de disminución y, en muchos casos, de desaparición definitiva (Larsen et al. 2012; UICN 2012). La pér-dida de biodiversidad está siendo particu-larmente importante en mares y océanos, los cuales albergan la mayor diversidad de ecosistemas y especies del planeta (Agardy 2005). Paradójicamente, el hecho de que los océanos sean patrimonio de todas las naciones crea un vacío legal que se erige como principal obstáculo para incrementar la protección de los mismos y de las espe-cies que viven en ellos (Hendriks et al. 2006; Norse 2010).

Una herramienta indispensable para la conservación de los ecosistemas marinos es la creación y manejo de Áreas Marinas Protegidas (AMPs), que se definen como ‘’áreas de los océanos designadas para me-jorar la conservación de los recursos mari-nos’’ (Lubchenco et al. 2003). En el ámbito concreto de la conservación de los verteb-rados marinos, las AMPs se han convertido en una herramienta fundamental para su protección (Abdulla et al. 2008; Botsford et al. 2009). Es evidente que no se puede pro-teger algo si se desconoce su distribución, por lo que la conservación de especies ma-rinas de gran movilidad exige un exhaus-tivo estudio previo de los patrones de uso espacial de su hábitat (Costello et al. 2010; Block et al. 2011), especialmente cuando el objetivo es proteger especies migratorias. Este es el caso de las tortugas marinas.

Las tortugas marinas son reptiles adaptados a la vida marina que habitan la Tierra des-

de el Cretácico (Hirayama 1998). En la ac-tualidad existen siete especies agrupadas en dos familias: familia Dermochelydae, con la tortuga laúd (Dermochelys coriacea) como única especie viviente, y la familia Cheloni-idae, con seis especies actuales: tortuga car-ey (Eretmochelys imbricata), tortuga bastarda (Lepidochelys kempii), tortuga golfina (Lepido-chelys olivacea), tortuga boba (Caretta caretta), tortuga verde (Chelonia mydas) y tortuga pla-na (Natator depressus) (Pritchard 1996; Spo-tila 2004). Las tortugas marinas dependen tanto de hábitats marinos como terrestres para su crecimiento y desarrollo. Estos háb-itats incluyen playas, arrecifes bentónicos y mar abierto. El ciclo de vida es similar en las siete especies, con pequeñas variaciones en la duración de las fases, presentando to-das ellas un comportamiento de anidación muy estereotipado (Miller 1997; Meylan y Meylan 1999; Ver figura 1.1 en página 51).

Las tortugas marinas son reconocidas como “especies clave” debido al impacto ecológico que tienen en el mantenimiento de la estructura y funcionamiento de los ecosistemas marinos (Bjorndal y Bolten 2003). En las últimas décadas, estas fun-ciones se están viendo alteradas debido a la drástica disminución de muchas pobla-ciones de tortugas en todo el mundo (Mc-Clenachan et al. 2006). La disminución de estas especies se debe a múltiples factores, que van desde la sobreexplotación huma-na (consumo durante siglos de sus huevos, carne y otros productos derivados de es-tos animales), hasta la captura accidental por barcos de pesca, el desarrollo costero, la contaminación o los efectos del cambio climático (Wallace et al. 2011).

La temperatura es un factor ambiental cru-cial en el desarrollo de las tortugas marinas,


ya que estas exhiben lo que se denomina determinación sexual dependiente de la temperatura (TSD, por sus siglas en in-glés), es decir, no existe una determinación genotípica del sexo, sino que la proporción de sexos primaria está influenciada por la temperatura experimentada por los huevos durante, aproximadamente, el tercio medio del periodo de incubación (Bull 1980; Yn-tema y Mrosovsky 1982). La temperatura pivote es aquella en la que el 50% de los neonatos que nacen son hembras y el 50% machos. Esta temperatura se sitúa en tor-no a los 29ºC para todas las especies; las temperaturas de incubación inferiores a la temperatura pivote producen más machos y las temperaturas superiores producen más hembras (Yntema y Mrosovsky 1982). Considerando este rasgo de las tortugas marinas, cualquier factor causante de va-riaciones en la temperatura de incubación afectaría a la proporción de sexos de los neonatos.

Alrededor del mundo se han establecido innumerables programas de conservación con el objetivo de proteger las poblaciones nidificantes de tortugas marinas amenaza-das. Una de las prácticas de protección más extendidas ha sido el traslado y reubicación de las nidadas amenazadas a corrales cerra-dos y protegidos, así como la incubación de los huevos en cajas de poliestireno (Brown et al. 2011; Maulany 2012), lo que puede tener resultados positivos en la conservación. Sin embargo, el traslado de las nidadas puede afectar negativamente al desarrollo embri-onario. No solo es posible que se produzca un sesgo de los neonatos hacia uno u otro sexo debido a variaciones en la tempera-tura, como hemos descrito anteriormente, sino que, además, tanto estas variaciones de

temperatura como el manejo y traslado de los huevos pueden afectar negativamente al desarrollo embrionario, al éxito de eclosión y al éxito de emergencia (Eckert y Eckert 1990; Rees y Margaritoulis 2004; Pintus et al. 2009; Santidrián Tomillo et al. 2009). A pesar de estos aspectos negativos, la falta de recursos para la adecuada conservación de las nidadas in situ hace que muchos proyec-tos de conservación, particularmente los establecidos en países en vías de desarrol-lo, no tengan otra opción que la de incubar los nidos artificialmente, bien en corrales ubicados en las mismas playas o en cajas (García et al. 2003; Brown et al. 2012). En aquellos casos en los que la incubación ar-tificial sea la única estrategia de protección posible es esencial estudiar sus efectos sobre el éxito de eclosión y las proporciones de sexo de los neonatos.

En la región del Caribe habitan seis espe-cies de tortugas marinas: la tortuga boba, la tortuga verde, la tortuga carey, la tortu-ga bastarda, la tortuga golfina y la tortu-ga laúd. Históricamente, la presencia de grandes colonias de anidación a lo largo de las costas caribeñas ha sido importante y, sin embargo, hoy en día, pocas pobla-ciones presentan una gran abundancia de tortugas nidificantes (McClenachan et al. 2006). Si tenemos en cuenta todas las es-pecies, aproximadamente la mitad del total de playas de anidación conocidas presentan menos de 25 rastros por año, lo que equiv-ale a menos de 10 hembras reproductiva-mente activas (Dow et al. 2007). En el Ca-ribe las tortugas marinas han sido objeto de explotación durante muchos siglos, lo que ha dado lugar a una reducción a gran esca-la en el número de individuos de sus pobla-ciones (Parsons 1962; Eckert 1995; Jackson

Resumen I 31

La desaparición de estas poblaciones po-dría tener un importante impacto a niv-el regional, por lo que su estudio debería considerarse un objetivo prioritario (Dow y Eckert 2011).

La República Dominicana (RD) está situ-ada en el extremo oriental de la isla de La Española. Es la segunda isla en tamaño dentro de las Antillas Mayores y uno de los puntos calientes de biodiversidad más im-portantes de América Central. En las zonas costeras de RD se ha descrito la anidación de cuatro especies de tortugas marinas: tor-tuga laúd, tortuga carey, tortuga verde y tortuga boba (Ottenwalder 1981). El país cuenta con un largo historial de consumo y comercio de estas especies, lo que ha constituido un importante recurso para las comunidades costeras (Ottenwalder 1981; Fleming 2001; Reuter y Allan 2006). La in-dustria de artesanía de concha de tortuga carey ha representado una grave amenaza para la supervivencia de esta especie, doc-umentándose durante décadas la venta de artículos de carey a turistas (Stam y Stam 1992; Domínguez y Villalba 1994; Feliz et al. 2008). De igual manera, ha sido gener-alizado el consumo de la carne, los huevos y la grasa de las tortugas verde y laúd. La pérdida de hábitat debido al desarrollo costero también constituye otra importante amenaza para los hábitats de anidación de las tortugas marinas en el país (Ottenwalder 1996; Geraldes 2003).

Las tortugas marinas están protegidas le-galmente y su comercio está prohibido en el país por la Ley de Pesca de 2005 CO-DOPESCA (Consejo Dominicano de Pesca y Agricultura). Sin embargo, pese a la exis-tencia de estas leyes y de áreas protegidas, la debilidad institucional se hace patente

1997; Meylan 1999; Bell et al. 2006). Las tortugas han sido muy valoradas tanto por sus huevos, como por su carne y caparazón (Fleming 2001; Bräutigam y Eckert 2006). A pesar de encontrarse estrictamente pro-tegidas mundialmente (todas estas especies se enumeran en el Apéndice I de CITES, que prohíbe su comercio a nivel internacio-nal), en algunos países del Caribe todavía se mantiene el comercio de productos de tortuga carey (Bräutigham y Eckert 2006). A estas amenazas directas se le suman los excepcionalmente altos niveles de desarrol-lo turístico que caracterizan esta región y que han provocado una degradación ambi-ental irreversible afectando a los hábitats de nidificación de las tortugas marinas (Dav-enport y Davenport 2006; Bell et al. 2007; Mathenge et al. 2012).

Recientemente se ha establecido que en el suroeste occidental del océano Atlán-tico la tortuga laúd presenta un estado de preocupación menor para su conservación (IUCN 2013). En la región del Caribe se considera que las poblaciones de esta espe-cie están controladas, y muchas de ellas en crecimiento (Dutton et al. 2005; McGowan et al. 2008; Wallace et al. 2011). Las pobla-ciones de tortuga carey y tortuga verde se han descrito como poblaciones robustas, pero bajo amenazas que, si no son elimina-das, podrían tener impactos negativos sobre el estado actual de las mismas a corto plazo (Wallace et al. 2011). A pesar de este apar-ente buen estado de las poblaciones de tor-tugas marinas en la región del Caribe, aún existen muchos vacíos de información so-bre el estado de conservación de pequeñas poblaciones dispersas, algunas de ellas muy diezmadas y amenazadas (Lagueux y Campbell 2005; Bräutigam y Eckert 2007).


biología reproductiva de las especies nidi-ficantes en las playas del Parque Nacional Jaragua (PNJ) y en las playas de isla Saona (Parque Nacional Del Este), y al mismo ti-empo evaluar los proyectos de incubación artificial establecidos durante el estudio en estas dos áreas para la protección de los nidos amenazados; 3) estudiar el compor-tamiento durante el periodo entre puestas de las hembras nidificantes y analizar el uso de sus áreas de anidación y alimentación en relación a las áreas protegidas de la región del Caribe .

Estos objetivos generales se concretan en los siguientes objetivos específicos:

- Identificar las principales colonias nid-ificantes de tortugas marinas en RD y describir los patrones espacio-tempo-rales de anidación en la actualidad, así como evaluar el posible impacto de las amenazas actuales sobre estas pobla-ciones nidificantes comparándolo con los estudios previos en el país.

- Evaluar el programa de incubación arti-ficial de nidos de tortuga laúd establecido en las playas del PNJ a través del estudio de los potenciales efectos de esta incu-bación artificial sobre el éxito de eclosión de los nidos y la proporción de sexos de los neonatos.

- Determinar el éxito reproductivo de Eretmochelys imbricata en la isla Saona a través del estudio del éxito de los nidos. Evaluar el programa de incubación ar-tificial establecido de la misma manera que para la tortuga laúd en el PNJ.

- Estudiar los factores que afectan al éxito de eclosión de los nidos de tortuga laúd incubados en las playas del oeste del PNJ

en el frecuente incumplimiento de la ley de medio ambiente y recursos naturales. Un ejemplo de ello es la enorme expansión del turismo en la costa, que ha llevado a una apropiación de parcelas de terreno protegi-do en zonas costeras (Heredia 2003).

Los últimos datos sobre el estado de conser-vación de las poblaciones nidificantes de tor-tugas marinas en RD son los trabajos lleva-dos a cabo por Ottenwalder (1981, 1983). Los resultados de estos trabajos mostraban un declive de las poblaciones nidificantes de tortugas en el país. Durante tres décadas no se ha realizado ningún otro estudio evalu-ando el estado de estas poblaciones, aunque han sido constantes las denuncias sobre los elevados niveles de explotación y comercio de productos procedentes de las mismas (Stam y Stam 1992; Domínguez y Villalba 1994). La falta de información actualizada sobre el estado de las poblaciones de tor-tugas marinas de RD ha sido resaltada de forma reiterada en diferentes evaluaciones sobre la situación de las tortugas marinas en la región del Caribe (McClenachan et al. 2006; Dow et al. 2007). La actualización del estado de conservación de las tortugas marinas en el país es necesaria no solo por la supervivencia de las tortugas, sino por el efecto que podría tener sobre la conser-vación de estas especies a nivel regional.

JUSTIFICACIóN Y OBJETIVOSEsta Tesis Doctoral tiene tres objetivos ge-nerales: 1) Llevar a cabo el primer estudio sistemático sobre el estado de conservación y amenazas actuales en las playas de ani-dación de tortugas marinas en las costas de RD en el Caribe Norte; 2) investigar la

Resumen I 33

km2, de los cuales 905 km2 son marinos. El parque abarca numerosos ecosistemas que destacan por el elevado número de especies endémicas presentes; estos ecosistemas in-cluyen bosques naturales, playas, costas ro-cosas, humedales, pastos marinos y arrecifes de coral. El parque cuenta con dos áreas de playas, el área de Bahía de las Águilas en el extremo occidental y las playas próximas a la laguna de Oviedo al este; ambas propor-cionan unas condiciones favorables para la anidación de las tortugas marinas laúd, carey y verde. El Parque Nacional del Este (PNDE, sureste de RD), se fundó el 16 de septiembre de 1975. En este parque se in-cluye la isla Saona que, con una superficie de 110 km2, es la mayor de las islas perte-necientes a RD. Las zonas neríticas adya-centes a las playas de la isla comprenden arrecifes de coral y praderas de pastos ma-rinos que proporcionan hábitats propicios para cientos de especies de plantas, aves, peces y otros animales marinos. Isla Saona es también el hogar de varias especies ani-males y vegetales endémicas, amenazadas o en peligro de extinción. Las playas de la isla son un hábitat favorable para la anidación de la tortuga carey y verde.

Muestreos de playasAntes de iniciar los muestreos de playa se revisó la información de los estudios llevados a cabo previamente por Otten-walder (1981), y se realizaron entrevistas a guardaparques y pescadores de las áreas de anidación descritas en tales estudios. Se realizaron recorridos a pie durante el día en las playas del PNJ y Saona durante cin-co temporadas de anidación consecutivas, 2006-2010. En las playas del este del PNJ, en las temporadas de 2006, 2007 y 2008 los

con el objetivo de determinar el valor re-productivo de estas playas para la espe-cie. Destacar el papel de la tortuga laúd en las playas amenazadas del PNJ como un vehículo para la conservación de los ecosistemas y la gestión en esta área pro-tegida.

- Investigar el comportamiento y las pref-erencias de hábitat de las hembras de tortuga carey nidificantes de isla Saona durante los periodos entre puestas. Es-tablecer las áreas preferentes de uso con respecto a los límites de las áreas prote-gidas en sus áreas de anidación y de ali-mentación.

MATERIALES Y MéTODOS

Área de estudioRepública Dominicana comparte con Haití la isla de La Española, que forma parte del archipiélago de las Antillas Mayores en la región del Caribe. El país tiene una exten-sión de 48.445 km2 y una población estima-da de unos 10 millones de personas. Cuen-ta con 1.576 km de costa. Tiene un clima tropical con una fuerte influencia maríti-ma, una temperatura promedio de 25 °C y rangos de precipitación que oscilan entre los 350 mm a los 2.500 mm. El Sistema Nacional de Áreas Protegidas cuenta con 86 unidades englobadas en seis categorías con diferentes niveles de protección (desde áreas de protección estricta hasta paisajes protegidos).

El presente estudio se llevó a cabo princi-palmente en las áreas protegidas del PNJ e isla Saona. El PNJ se encuentra situado en el suroeste de RD, comprende 1.374


datos sobre la fecha y la hora de puesta, ubicación del nido en relación a la línea de marea alta (distancia), ubicación a lo lar-go de la playa (sector) y zona (vegetación, borde, arena). Se utilizó una cinta métrica flexible para tomar las medidas en curvo de cada tortuga: longitud curva de caparazón (LCC) y ancho curvo de caparazón máximo (ACCmax). Las tortugas laúd se marcaron con placas de metal Inconel grandes (Mar-ca nacional y Tag Co., Newport, EE.UU.) entre la cola y aletas traseras. Las tortugas carey se marcaron con marcas de metal In-conel pequeñas en el borde de las dos aletas delanteras.

Una vez que la tortuga abandonó la playa se procedió al manejo de los huevos. El des-tino de las nidadas se decidió en función del riesgo de inundación del nido por la marea, de la presencia de depredadores y/o de la distancia a campamentos pesqueros y po-blaciones. Si no se detectaron amenazas para los huevos, la nidada se dejó en la pla-ya para su incubación en condiciones na-turales (in situ), camuflando el rastro de la tortuga para evitar su detección. En aquel-los casos en los que existía riesgo de pérdi-da de los huevos por alguna amenaza, se procedió a la excavación del nido para el traslado de los huevos a las casetas de los guarda parques, donde se incubaron en cajas de poliuretano (dimensiones: 30 cm ancho x 50 cm largo x 32 cm profundidad). El manejo y traslado de huevos se hizo con extremo cuidado, siguiendo la metodología descrita, para minimizar la mortalidad de embriones (Abella et al. 2007). Los latera-les y parte inferior de las cajas se cubrieron con arena procedente del nido original para evitar el contacto de los huevos con las paredes. Igualmente, una vez en la caja, los

recorridos se realizaron una vez por sem-ana. A partir de 2009, los guarda parques recorrieron diariamente esta playas. Las playas del oeste del PNJ se recorrieron en-tre 3 y 4 veces por semana. En la isla Saona, las playas se recorrieron al menos una vez por semana durante todo el año; a partir de 2008 estos recorridos se incrementaron a 3-4 por semana durante el periodo de máxima anidación (de junio a noviembre).

Durante estos recorridos se registró infor-mación de todos los rastros de tortugas encontrados, identificándose cada especie mediante el estudio de características pro-pias del rastro de las mismas (Pritchard y Mortimer 1999). Para determinar el núme-ro de nidos en playa se diferenció entre rastros de tortugas con anidación exitosa (depositaron huevos) de aquellos rastros de tortuga sin anidación exitosa. Debido a las limitaciones de financiación sólo fue posible llevar a cabo recorridos nocturnos en isla Saona y en las playas occidentales del PNJ.

La verificación de presencia de anidación en otras áreas de RD se llevó a cabo me-diante visitas, recorridos esporádicos y en-trevistas a pescadores y habitantes de las comunidades locales durante el periodo 2006-2010. Este trabajo se realizó en 11 zo-nas de la costa norte en 2006, 2007 y 2008, y en seis playas de las costas este y sur del país en 2008, 2009 y 2010. Los informes sobre otros eventos de anidación, en partic-ular aquellos procedentes de hoteles o com-plejos turísticos situados frente a la playa, también fueron registrados.

Toma de datos y traslado de nidadasLos nidos encontrados se posicionaron me-diante coordenadas GPS. Se recogieron

Resumen I 35

La determinación de la proporción de sexos se hizo mediante métodos indirectos, utilizando el periodo de incubación (PI). Para cada nido, se calculó el PI (definido como el número de días entre la puesta de huevos y la primera eclosión de neonatos). Dado que no existe ningún estudio previo sobre el PI pivote (PI en el que nacen 50% machos y 50% hembras) para Saona, utili-zamos el valor estimado para el área de ani-dación de carey más próxima, Isla de Mona (Mrosovsky et al. 2009), y en el caso de la tortuga laúd se utilizó la curva que relacio-na PI y proporción de sexos estimada para la población de Suriname (Godfrey 1997). Se extrapolaron sobre la curva los valores de PI registrados para cada nido, obtenien-do de esta manera la proporción de sexos correspondiente a cada PI.

Telemetría por satélite Se colocaron transmisores vía satélite a tor-tugas carey interceptadas en el momento de desove en las playas de isla Saona, tras haber realizado las puestas, siguiendo la metodología de Godley et al. (2003).

Las posiciones de cada tortuga se deter-minaron a través del sistema ARGOS. Este sistema de localización y recolección de da-tos por satélite asigna un valor de precisión a las posiciones estimadas (location class [LC] 3, 2, 1, 0, A, B, Z). Los datos de teledetec-ción se descargaron y se filtraron mediante la herramienta de análisis del seguimiento por satélite (STAT) (Coyne y Godley 2005). En este estudio las LCs seleccionadas para la estimación de las áreas de uso fueron 3, 2, 1, A y B (señaladas en trabajos previos como las más precisas en estudios de carey; Gaos et al. 2012).

huevos se cubrieron también con arena del nido original. Cada caja fue etiquetada con un código, indicando el nombre de la playa, la fecha de puesta y el número de huevos incubados. Las cajas fueron revisadas dia-riamente a lo largo de todo el periodo de incubación y las tapas se abrieron durante dos o tres horas al día para permitir la cir-culación del aire.

Éxito de eclosión y éxito de emergenciaEn los nidos incubados artificialmente, se procedió al estudio de los huevos siempre 48 horas después de la emergencia del último neonato. Las crías fueron liberadas cuando entraron en estado de frenesí y, siempre que fue posible, las liberaciones se realizaron en las playas de procedencia de los nidos. An-tes de su liberación, se escogieron aleatori-amente 20 neonatos y se tomaron medidas de la longitud recta del caparazón (LRC) con un pie de rey con una precisión de 0,1 cm. Estos 20 neonatos se pesaron con una balanza electrónica con una precisión de 0,1 g. Los nidos que se dejaron incubando in situ se controlaron durante los muestreos de playa; cuando se observaron rastros de neonatos saliendo del nido se llevó a cabo la excavación y estudio del contenido. La clasificación del contenido de los nidos y el cálculo del éxito de eclosión y éxito de emergencia se detallan en las páginas 81, 82 y 83 del capítulo 3.

Estimación de la proporción de sexos de los neonatosDebido al estado de conservación de la po-blación nidificante de Saona no se consid-eró la opción de sacrificar neonatos para la determinación del sexo mediante histología.


y tortuga verde (media = 9,2 ± 6,2, ran-go: 1-15, n = 37) de todo el país, mientras que de tortuga laúd se registraron 22 nidos en el mismo periodo (media = 5,5 ± 4,8, rango: 1-11). La anidación de las diferentes especies registrada en el resto de playas del país es muy reducida. El número de nidos por playa se presenta en la figura 4.2 de la página 96. El periodo de anidación de la tortuga laúd en el PNJ se extiende desde marzo hasta agosto con la mayoría de nidos registrada entre los meses de abril y junio (89,6% de los nidos). En la isla Saona, la tortuga carey fue observada anidando a lo largo de todo el año con la mayoría de ni-dos registrada en los meses comprendidos entre junio y noviembre. La tortuga verde anida en Saona de julio a noviembre, con el máximo de nidos registrados en el mes de agosto. Tanto en el PNJ como en Sa-ona, todas las tortugas detectadas en los muestreos realizados durante el periodo de estudio fueron identificadas con una mar-ca metálica numerada. La media de LCC de la tortuga laúd en el PNJ es de 147,4 ± 8,7 cm, n = 13. En Saona se midió el LCC de un total de 19 hembras de carey (media 87,2 ± 4,7 cm). La talla media de las tortu-gas carey que anidan en RD es similar a la registrada en otras áreas de anidación del Caribe. En el caso de la tortuga laúd, la tal-la media registrada en el PNJ es inferior a la registrada en la mayoría de las poblaciones del mundo. Este dato podría interpretarse como presencia de una elevada proporción de hembras neófitas, lo que indicaría una posible recuperación de la población.

La depredación de huevos por humanos se identificó como la principal amenaza so-bre los nidos de tortugas marinas en playas de anidación de RD. En el PNJ, los nidos

Para estimar el área de actividad de las hembras nidificantes dentro de las zonas entre puestas y las zonas de alimentación se utilizó la herramienta de mínimo polígono convexo (MPC). La intensidad de uso de las áreas de actividad se estableció mediante el estimador de densidad kernel (EDK) (Wor-ton 1989). Las distribuciones de densidad fueron representadas en los mapas utilizan-do los contornos de distribución del 50% y 90% (Hooge et al. 2000).

La ubicación de las tortugas con respecto a las áreas marinas protegidas (AMP), se anal-izó superponiendo los límites de estas áreas en todos los mapas resultantes. Los límites de las AMPs fueron descargados de la base de datos mundial de áreas protegidas (www.wdpa.org). Para llevar a cabo estos análisis se utilizó el programa de sistemas de infor-mación geográfica ArcGIS v.10.0 (ESRI 2010).

RESULTADOS Y DISCUSIóNLos resultados de este estudio confirman la presencia de tres especies de tortugas mari-nas anidando actualmente en RD: la tortu-ga laúd, la tortuga carey y la tortuga verde. El PNJ presentó el mayor número de nidos por año de tortuga laúd (media = 126,4 ± 74,1, rango: 17-210) con un total de 632 ni-dos contabilizados durante el periodo 2006-2010. El número total de nidos de tortuga carey registrados en el mismo periodo en el PNJ fue de 73 (media 14,6 ± 6,7, rango: 7-22). En los cinco años de estudio solo se registró un nido de tortuga verde en esta área. En las playas de isla Saona se registró el mayor número de nidos de tortuga carey (media 100 ± 8,4, rango: 93-111, n = 400)

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Resumen I 37

paración con los nidos incubados en playas.

Durante el periodo de estudio se estudiar-on un total de 109 nidos de tortuga laúd (puestos en las temporadas 2008 y 2009): 35 incubados artificialmente en las playas del este (PE), 31 incubados artificialmente en las playas occidentales (PO) y 43 incuba-dos in situ también en las playas occidentales del parque. Los resultados obtenidos mues-tran que el método de incubación influyó significativamente en el éxito de eclosión de las nidadas (GLMM: χ2 = 76,7, p < 0,001). El éxito de eclosión fue significativamente más alto en los nidos incubados in situ que en los incubados artificialmente, tanto en PO como en PE. El método de incubación tuvo también un efecto sobre el número de embriones muertos en estado de desarrollo temprano (GLMM: χ2 = 16,1, p < 0,05). Los nidos incubados artificialmente en PO tenían más embriones muertos en estado temprano que los nidos incubados in situ y los incubados artificialmente en PE. El número de embriones muertos en estado tardío también se vio afectado por el méto-do de incubación (GLMM: χ2 = 67,4, p < 0,001). Los nidos incubados artificialmente en PE y PO presentaron un número signif-icativamente mayor de embriones muertos en estado tardío que los nidos incubados in situ. Igualmente, el método de incubación afectó a la duración de la misma (GLMM: χ2 = 16,0, p < 0,01). Las temperaturas de incubación fueron significativamente inferi-ores en los nidos incubados artificialmente en PE en comparación con los incubados artificialmente en PO (prueba t de Welch, t22,9 = -5,66, p < 0,001). El lugar de incu-bación de los nidos incubados artificial-mente en PO está situado en una zona elevada de matorral árido, por lo que los ni-

de las playas situadas en el extremo orien-tal vienen sufriendo los niveles más altos de depredación. El número total de nidos depredados en playa por año de cada una de las tres especies estudiadas se muestra en la tabla 4.2 de la página 99. Tanto en el PNJ como en Saona se encontraron restos de adultos de tortuga carey y tortuga verde presumiblemente depredados para con-sumo humano.

Los resultados de este estudio muestran que la anidación de tortugas marinas en el país está restringida a las áreas protegidas y poco desarrolladas del PNJ e isla Saona. Cuando comparamos los resultados obteni-dos en este trabajo con los obtenidos en trabajos previos llevados a cabo en el país, se observa un declive de las poblaciones nidificantes de tortugas marinas. Durante el periodo de estudio se observó un manejo inadecuado de las áreas protegidas por par-te de las autoridades locales, registrándose un alto grado de depredación y consumo tanto de huevos como de carne de tortugas adultas por habitantes locales. Estos resul-tados ponen de relieve la necesidad de una mejor protección y manejo de estas áreas.

Como medida frente a la elevada depre-dación humana se continuó un programa de protección de nidos establecido desde 1974, en el que los huevos eran trasladados desde la playa para su incubación en cajas. Este sistema de protección de nidos de tor-tuga laúd se ha realizado voluntariamente durante treinta y ocho años de forma es-porádica bajo escaso control científico. En el presente estudio hemos tratado de eval-uar la eficacia de esta medida de conser-vación investigando si la incubación artifi-cial puede alterar el éxito de eclosión y las proporciones de sexo resultantes en com-


cajas como medida urgente para proteger los nidos e incrementar el reclutamiento de neonatos al mar.

Durante los cuatro años de estudio se reg-istraron 400 nidos de tortuga carey. En to-tal se incubaron artificialmente 163 nidos, otros 146 se encontraron ya depredados por humanos y 91 se dejaron incubando in situ. La figura 6.3 de la página 140 muestra la variación anual en estos porcentajes. En todos los nidos depredados por humanos se depredó el 100% de los huevos.

En estos cuatro años de estudio se liberaron al mar un total de 12340 neonatos. Las biometrías de los neonatos liberados sugi-eren que éstos son más pequeños que los de otras áreas de puesta del Caribe (2008: LCC = 3,8 ± 0,1 cm, rango: 3,2-4,2, n = 480); 2009: 3,8 ± 0,2 cm, rango: 3,0-4,0, rango: 12,6-18,2, n = 160). No se detec-taron efectos significativos de la incubación artificial sobre el éxito de eclosión entre años (ANOVA: F(3,157)= 1,45, P = 0,383), tipo de incubación (F(1,157) = 0,7, P = 0,45); o su interacción (F(3,157) = 0,52, P = 0,669). Igualmente no hubo diferencias significa-tivas en el éxito de emergencia entre años (ANOVA: F(3,157) = 1,1, P = 0,469), el tipo de incubación (F(1,157) = 1,28, P = 0,327); o la interacción de ambos factores (F(3,157) = 0,68, P = 0,563), comparado con los nidos incubados in situ. Se obtuvo el periodo de incubación (PI) de un total de 84 de los ni-dos incubados artificialmente y de 24 de los nidos incubados in situ. Se observaron peri-odos de incubación más largos en los nidos incubados en cajas que en los incubados in situ. Por otro lado, en 21 de los 22 nidos en los que se registró la temperatura de incu-bación durante el periodo termosensible, ésta fue menor que la temperatura pivote

dos alcanzaron altas temperaturas durante la incubación. Como resultado, la duración de la incubación y las proporciones de sexo de los neonatos incubados en esta ubicación fueron similares a las obtenidas en los nidos incubados in situ. Los nidos incubados ar-tificialmente en PE tuvieron duraciones de incubación significativamente más largas que los nidos incubados artificialmente en PO y los incubados in situ. Esto se tradujo en porcentajes de hembras muy bajos en PE, probablemente debido a que el lugar de incubación estaba ubicado en una zona de mayor humedad (bosque junto a la lagu-na de Oviedo). Dado que en estas playas no se incubaron nidos in situ, no se pudo com-probar las diferencias en la proporción de sexos entre nidos incubados en condiciones naturales y nidos incubados artificialmente.

En las playas occidentales del PNJ la in-cubación de los nidos in situ parece posible mejorando sensiblemente la protección de playas. Sin embargo, en las playas del este del parque esta opción no es viable actual-mente debido a la intensa depredación hu-mana que existe. En las playas orientales del PNJ la incubación artificial parece la única opción para la incubación con éxito de los huevos, aunque, teniendo en cuen-ta los bajos porcentajes de hembras que se obtienen en los nidos incubados artificial-mente, es necesario implantar medidas de conservación que aseguren que la propor-ción de sexos de los neonatos sea similar a la producida en los nidos incubados in situ.

Al igual que ocurre en las playas del PNJ, la colonia reproductora de tortuga carey de isla Saona se encuentra seriamente amenazada principalmente por la depredación huma-na de huevos. En 2007 se inició en Saona un programa de incubación de huevos en

Resumen I 39

los nidos depositados al comienzo de la temporada de puesta presentaron éxitos de eclosión significativamente superiores a los incubados al final de la temporada. Los ni-dos de tortuga laúd incubados en las playas occidentales del PNJ presentan un éxito de eclosión inusualmente alto (75,2 %) para esta especie. Los resultados de este estudio son especialmente relevantes en relación a la playa de La Cueva (figura 7.1 en página 158). Este sector recibe el 20% del total de nidos de tortuga laúd en las playas occiden-tales del PNJ, nidos que además registraron los niveles más altos de éxito de eclosión. Sin embargo, este sector está menos prote-gido, ya que se encuentra en la zona de am-ortiguamiento del PNJ, fuera de los límites del área protegida. Teniendo en cuenta los excepcionales niveles de éxito de eclosión y las amenazas actuales y potenciales que afectan a las playas de anidación de tortuga laúd, se requiere un esfuerzo adicional en la regulación, gestión y redefinición de esta área protegida.

Realizar una protección eficaz de especies migratorias requiere conocer los patrones espaciales y temporales de uso de hábitat de las mismas en las diferentes etapas de su ciclo vital. De estos estudios depende en gran medida la eficacia en la gestión y conservación de las áreas marinas protegi-das (AMPs) (Maxwell et al. 2011; Scott et al. 2012). En este trabajo se estudió el com-portamiento entre puestas y post-puesta de las tortugas carey nidificantes en Saona. Se colocaron transmisores de satélite a un total de 9 hembras nidificantes: 8 en Saona, en 2008 y 2009; y 1 en el PNJ en 2009.

Durante los periodos entre puestas las hem-bras nidificantes de tortuga carey en Saona permanecieron en las aguas adyacentes a

estimada en estudios realizados en el labo-ratorio recogidos en la literatura (Mrosovsky et al. 2009). Los resultados de periodos y temperatura de incubación sugieren un ses-go hacia la producción de machos en estos nidos incubados artificialmente.

Los actuales niveles de depredación de hue-vos por humanos no son sostenibles para la conservación a largo plazo de esta colonia reproductora. Sin embargo, los bajos por-centajes de hembras obtenidos en los nidos incubados en cajas muestran la necesidad de cambios urgentes en la estrategia de conservación adoptada.

En los últimos treinta años, RD ha sufri-do un desmesurado desarrollo de lo que se conoce como “turismo de masas”, especial-mente en las costas norte y este del país. Re-cientemente, este modelo de turismo, junto con la explotación de minas de bauxita, ha sido propuesto como vía para el desar-rollo de la región suroeste del país. Dentro de estos planes de desarrollo se incluyen proyectos para la construcción de com-plejos turísticos dentro del área protegida del PNJ. En el presente estudio se analizó el éxito de eclosión de 64 nidos de tortuga laúd puestos durante tres temporadas de anidación (2007, 2008 y 2009). Los resul-tados muestran que el éxito de eclosión de los nidos estudiados parece estar determi-nado por tres factores: el sector de la playa donde se incubaron, la duración del PI y la fecha de puesta. El éxito de eclosión de los nidos incubados en el sector de La Cueva fue significativamente más alto que el de los nidos incubados en el sector de Bahía de las Águilas. El PI afectó al éxito de eclosión, los nidos con PI más largos presentaron éxitos de eclosión más bajos. La fecha de puesta también afectó al éxito de eclosión:


mentación. Una de las tortugas que perma-neció durante su periodo de alimentación en aguas dominicanas, se mantuvo dentro de los límites del PNJ, mientras que el área de alimentación de la segunda se situó en aguas adyacentes a Bahía de las Calderas (costa sur de RD), fuera de cualquier área marina protegida. La última tortuga ras-treada viajó al noroeste de las Bahamas, donde su área de actividad se mantuvo fuera de áreas protegidas.

La información obtenida pone de relieve la importancia de las áreas protegidas de RD para la tortuga carey, tanto como hábitat entre puestas como hábitat de alimentación, y muestra la necesidad de hacer cumplir la legislación vigente de las áreas protegidas del país. El presente estudio también con-firma que las aguas de Nicaragua y Hondu-ras son áreas de alimentación excepcional-mente importantes para la tortuga carey en el Caribe, y hace patente la vulnerabilidad de las tortugas en estas aguas.

sus playas de anidación. Los MPCs obteni-dos muestran que las áreas ocupadas por las tortugas entre puestas variaron entre 51,0 y 644,0 km2 (media = 254,3 ± 173,6). Las áreas de mayor intensidad de uso de las tortugas nidificantes en Saona se localiza-ron en el extremo oriental de la isla. Estas áreas se caracterizaron por ser aguas poco profundas, situadas principalmente dentro de la isobata de los 200 m de profundidad, y estar asociadas a arrecifes de coral. Las áreas de distribución se solaparon y su ubi-cación y extensión fueron similares entre tortugas y entre años. Nuestros resultados confirman el comportamiento de tortugas carey nidificantes durante el periodo entre puestas descrito previamente en otras áreas del Caribe (van Dam et al. 2008; Marcoval-di et al. 2012; Walcott et al. 2012). Dentro de este hábitat entre puestas descrito, los EDKs del 90% y del 50% tuvieron una ex-tensión estimada de 81,7 km2 y 32,2 km2 respectivamente. Esta área de uso común se localizó dentro de los límites del área marina del PNDE.

Las tortugas mostraron una dicotomía en sus estrategias migratorias: algunas perma-necieron relativamente cerca de sus playas de anidación, en aguas de RD, y otras mi-graron a zonas de alimentación internacio-nales. La mayoría de las tortugas migraron a áreas de alimentación situadas en aguas de Nicaragua y Honduras (n = 5). En es-tas aguas, con la excepción de dos tortugas que estuvieron el 2,2% y el 91,5% de los días dentro de áreas marinas protegidas (Reserva de la Biosfera Seaflower y Cay-os Misquitos respectivamente), el resto de tortugas permanecieron durante todo el periodo de alimentación fuera de cualquier área protegida durante sus periodos de ali-

Resumen I 41

del mundo. De este dato podría interpre-tarse la presencia de una elevada propor-ción de hembras neófitas, lo que indicaría una posible recuperación de la población. En el PNJ también se registró anidación es-porádica de tortugas carey y verde.

4. La isla Saona acoge las playas más im-portantes del país para la anidación de tor-tuga carey (media anual = 100 ± 8,4 nidos, rango 93-111, n = 400) y tortuga verde (media 9,2 ± 6,2, rango 1-15, n = 37). La tortuga carey anida en la isla durante todo el año, siendo el periodo de máxima ani-dación el comprendido entre los meses de junio y noviembre. El número estimado de hembras de carey anidando por año varió entre 21 y 25. La talla media de las tortugas carey que anidan en RD (LCC = 87,2 cm) es similar a la registrada en otras áreas de anidación del Caribe.

5. Las principales amenazas para la con-servación de estas poblaciones nidificantes de tortugas marinas son el elevado nivel de depredación de huevos para consumo hu-mano y el desarrollo costero. Como medida de protección frente a la depredación hu-mana se establecieron programas de incu-bación artificial de nidos.

6. En el PNJ la incubación artificial se llevó a cabo por separado en las playas orientales y occidentales del parque. Se compararon el éxito de eclosión y la proporción de sexos de los neonatos entre los nidos incubados artificialmente y los nidos incubados in situ. Los resultados mostraron que en los nidos artificiales incubados en el oeste el éxito de eclosión fue muy bajo, mientras que en los incubados artificialmente en el este los periodos de incubación indicaron un sesgo hacia a la producción de machos.

CONCLUSIONESLa presente tesis estudia el estado de conser-vación de las tortugas marinas nidificantes de RD. Las conclusiones del presente estu-dio son las siguientes:

1. En este estudio se ha realizado la pri-mera evaluación detallada del estado de conservación de las poblaciones de las tor-tugas marinas nidificantes en RD a partir de muestreos realizados durante el periodo 2006-2010. Se ha constatado la anidación de tres especies de tortugas marinas: tortu-ga laúd, tortuga carey y tortuga verde; sin embargo, durante el periodo de estudio no se detectó la presencia de tortuga boba, cuya anidación había sido descrita en el país en estudios previos.

2. Actualmente, la anidación de tortugas marinas se concentra en las áreas protegi-das del PNJ e isla Saona (PNDE). La ani-dación en playas fuera de estas áreas es esporádica. Los resultados de este estudio sugieren que se ha producido una gran re-ducción en la abundancia de las cuatro es-pecies de tortugas marinas en el país desde la década de 1980.

3. El PNJ alberga, con un total de 632 ni-dos registrados, el mayor número de nidos por año de tortuga laúd de todas las áreas del país estudiadas (media = 126,4 ± 74,1, rango 17-210). El periodo de anidación se extiende desde marzo hasta agosto, regis-trándose el mayor número de nidos en los meses de abril a junio. El número estimado de hembras de laúd anidando por año varió de 3 a 40. El tamaño de las tortugas laúd en RD (LCC medio = 147,4 cm) es inferior al registrado en la mayoría de las poblaciones


or excepcional de éxito de eclosión y las amenazas actuales y potenciales que afec-tan a las playas de anidación de la tortuga laúd, se requiere un esfuerzo adicional en la regulación y gestión de esta área protegida.

11. El sector de La Cueva alberga el 20% del total de nidos depositados en las playas occidentales del PNJ, presentando además los niveles más altos de éxito de eclosión. Este sector se encuentra fuera de protección en la zona de amortiguamiento, fuera de los límites del parque, por lo que se recomien-da su inclusión dentro del PNJ.

12. A través del seguimiento mediante telemetría por satélite de tortugas carey nid-ificantes en Saona, se constató que, durante el periodo entre puestas, estas tortugas se mantuvieron en las aguas territoriales RD, en su mayoría sobre la plataforma conti-nental (<200 m), en zonas caracterizadas por aguas relativamente poco profundas y cercanas a las playas de anidación.

13. El área de uso común estimada para las tortugas carey nidificantes en Saona duran-te los periodos entre puestas se situó den-tro de los límites del PNDE. Las áreas de uso individual se localizaron en el extremo oriental de la isla y mostraron similar ubi-cación y extensión entre tortugas y años. Se recomienda aumentar los esfuerzos para mitigar la pesca ilegal y para restringir el tráfico de embarcaciones en estas aguas, a fin de proteger a las tortugas en su hábitat entre puestas.

14. Durante el periodo de alimentación, el 78,0% de las posiciones obtenidas se en-contraron fuera de cualquier área marina protegida, tanto en aguas de RD como en aguas internacionales de Bahamas, Nicara-

7. La incubación de los nidos fuera de la playa es la única opción de conservación vi-able en las playas orientales del PNJ, dada la elevada depredación humana (100% de nidos y huevos), pero es necesario evitar la posible desviación de la proporción de sexos natural que ocasiona la incubación artificial. En las playas occidentales la con-servación de los nidos in situ parece posible a través de la mejora en el manejo y protec-ción de las playas.

8. El programa de incubación artificial lleva-do a cabo en Saona permitió la liberación de más de 12000 neonatos de tortuga carey. No se encontraron diferencias en el éxito de eclosión y emergencia entre los nidos incu-badas in situ y los incubados artificialmente. Sin embargo, las bajas temperaturas y lar-gos periodos de incubación registrados en los nidos incubados artificialmente sugieren un sesgo hacia la producción de machos. A pesar de que la incubación artificial es efi-caz en cuanto a la producción de neonatos, la escasa producción de hembras pone de manifiesto la necesidad de mejorar esta me-dida de protección.

9. Se estudiaron los factores que afectan al éxito de eclosión de los nidos de tortu-ga laúd incubados en las playas occidental-es del PNJ. El sector de la playa donde se incubaron, la duración del periodo de in-cubación, la fecha de puesta y el número de huevos incubados parecen ser los prin-cipales factores que determinan el éxito de eclosión.

10. Las nidadas de tortuga laúd de las pla-yas occidentales del PNJ presentaron un inusual elevado éxito de eclosión para esta especie (75,2 %) comparado con otras po-blaciones del caribe (~ 50%). Dado el val-

Resumen I 43

gua y Honduras. Nuestros resultados desta-can la importancia de diferentes áreas pro-tegidas de RD como áreas de alimentación para la tortuga carey, mostrando la necesi-dad de hacer cumplir la legislación exis-tente y de expandir algunas de las áreas protegidas en el país.

15. El presente estudio también corrobora que las aguas de Nicaragua y Honduras son áreas de alimentación de excepcional importancia para las tortugas carey que anidan en la región del Caribe, y muestra la vulnerabilidad de las tortugas en dichas aguas.

16. A pesar de la importancia de las áreas marinas y costeras protegidas, no sólo para las tortugas marinas nidificantes, sino tam-bién para la biodiversidad marina en el país, la gestión de estas áreas es limitada de-bido a la falta de recursos. La recuperación de las poblaciones nidificantes de tortugas en el país depende de una gestión compe-tente de las áreas protegidas, así como de la aplicación efectiva de las leyes que prohí-ben el consumo y el comercio de cualquier producto de tortuga marina.

1. CHAPTER I: GENERAL INTRODUCTION©

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ChApTEr I:GENERAL INTRODUCTION

Chapter I: General Introduction I 47

1. 1. BIODIVERSITY LOSS AND CONSERVATION

The maintenance of biodiversity (genetic, population, species and ecosystem diversi-ty) is considered to be one of the highest conservation priorities of our time (Brooks et al. 2006; Larsen et al. 2012). If we revise the biodiversity studies conducted in the last decades we can confirm that the entities that constitute our biosphere are individu-ally and collectively highly endangered or outright lost, with many species declining to critically low levels and with significant numbers becoming extinct (IUCN 2011). In the last four decades, there has been a decrease in individual populations of many species and a large number of habitats have suffered the loss of their original con-ditions (Brooks et al. 2006; Laurance 2007; Butchart et al. 2010). This biodiversity loss has become one of the most pressing crises, which has led to a growing global concern about the status of the biological resourc-es on which human life depends (ONU 2008; CBD 2010). Habitat loss and frag-mentation, overexploitation, the impact of invasive alien species, wildlife trade, pollu-tion and climate change have been identi-fied as the main factors causing the loss of species and ecosystems in both terrestrial and marine environments (Jackson et al. 2001; Dulvy et al. 2003; Lotze et al. 2006; Botkin et al. 2007; Stork et al. 2009; Wlo-darska-Kowalczuk et al. 2010).

Overexploitation means harvesting species from the wild at rates faster than natural populations can recover. This threat is asso-ciated with the extinction of many endan-gered vertebrates both in land and marine

ecosystems (Butman and Carlton 1995; Rosser and Mainka 2002; McClenachan et al. 2006). Overexploitation presents itself in many forms: exhausting a species used for food, clothing, or medicinal therapies (Alves and Rosa 2007). Sometimes organ-isms are harvested or captured for purposes other than food and many species are trad-ed to be pets, souvenirs, or trophies (Spur-geon 1992; Fleming 2001).

Although one of the main current challeng-es in conservation is to increase efforts to stop global biodiversity loss (Hoffmann et al. 2010), recent studies show that the rate of this loss does not seem to be slowing (Butchart et al. 2010). Currently, there is no evidence of mass global extinctions; howev-er, evidence does demonstrate high levels of species lost at the local and regional scale, with corresponding negative effects upon their ecosystems (Stork 2010). This is par-ticularly important in marine ecosystems, where human activities are driving many species to their ecological extinction (Jack-son et al. 2001).

1. 2. WHAT IS CONSERVATION?Conservation is defined (in an anthropo-centric and utilitarian way) as “the man-agement of human use of the biosphere so that it may yield the greatest sustainable benefit to present generations while maintaining its potential to meet the needs and aspirations of future generations” (IUCN 1980). In response to the biodiver-sity loss, a new scientific discipline called Conservation Biology was consolidated in the eighties of the 20th century (Soulé and Wilcox 1980). The main objectives of this discipline are to evaluate the factors


and reasons of the decline of biodiversity, develop tools and useful measures to pre-vent its loss, and provide answers to specific questions that can be applied to manage-ment decisions (Wilson 1992; Hendriks et al. 2006). The design of protected areas, establishment of breeding programs to prevent loss of genetic variability in small populations, or reconcile the conservation needs with needs of local communities are some of the tools used for the preservation of biodiversity (Primack 1993; Robinson 2006; Griffiths and Pavajeau 2008). The challenges of conservation biodiversity lie in the realm of organizational behaviours, human values, policy-making processes, le-gal structures, communication flows, public education and agency culture.

1. 3. MARINE BIODIVERSITY CONSERVATION

Most of the Earth’s surface (70.8%: 362 mil-lion km2) is covered by oceans and seas, which are home to the planet’s greatest diversity of ecosystems and species (Agardy 2005). Cur-rently, about two-thirds of the world’s popula-tion lives within 60 km of the coast, with many people relying directly on resources provided by the oceans. This high population density in coastal regions is threatening the ecological integrity of the marine environment, which is becoming increasingly degraded due to the compounding effects of various human ac-tivities (Rogers and Laffoley 2011). Few cor-ners of the world’s oceans remain unaffected (Halpern et al. 2007).

Marine ecosystems are undergoing a great loss of diversity due to direct causes such

as overexploitation by fisheries, coastal de-velopment, increased pollution and habitat destruction, as well as indirect causes such as climate change (Paramo et al. 2009; Cheung et al. 2010; Polidoro et al. 2010). Overfishing is the primary driver of biodi-versity loss in marine systems (Hoffmann et al. 2010). Throughout the 20th century, fishing fleets were transformed from small-scale operators to large industrial ships that employ the use of trawlers, netting, and other high capacity tools. The world’s cur-rent fishing capacity is estimated to be up to 2.5 times more than what is needed to land a sustainable yield (Sumaila et al. 2012). In addition, bycatch (non-target catch) has contributed to the decline of endangered marine species, including marine turtles, cetaceans, sharks, birds, and lots of other animals over short timescales (i.e. decades) (Lewison et al. 2004; Wallace et al. 2013).

In coastal zones several activities pose threats to marine biodiversity, most of which are a direct result of the increase of human population worldwide and demo-graphic trends (Gray 1997). The escalating population density in the coastal areas has intensified pressure on the use of marine and coastal resources provoking habitat degradation, fragmentation and destruc-tion (Nicholls et al. 2007). In addition, the growth of large coastal cities and industri-alisation has brought about a rise in pollu-tion levels - another major threat to marine biodiversity. Human activities are mainly to blame for the entry of chemicals, particles, industrial waste, agricultural and residen-tial waste, plastics and other solid materi-als into the ocean (Derraik 2002; Cole et al. 2011). Habitat loss represents another alarming threat in marine ecosystems. For


1. 4. COASTAL AND MARINE PROTECTED AREAS

Marine Protected Areas (MPAs) are de-fined as “areas of the ocean designated to enhance conservation of marine resources” (Lubchenco et al. 2003). MPAs have gained increasing popularity worldwide as tools for biodiversity conservation and ecosystem management. They also help to raise public awareness (Abdulla et al. 2008; Botsford et al. 2009; Guidetti et al. 2013). MPAs have been established with diverse objectives, ranging from conservation of biodiversity, protection of a particular species, groups of species or critical areas (Allison et al. 2003; Micheli et al. 2008), for the prevention of overfishing (Gell and Roberts 2002), and even for the enhancement of local fisheries when MPAs act as suppliers of adult fishes to adjacent non-protected sites (Roberts et al. 2001; Harmelin-Vivien et al. 2008).

Globally, there are over 4000 separate areas designated as MPAs (Wood 2007); howev-er, the actual level of protection, manage-ment and regulation within MPAs varies considerably according to its designation (Lubchenco et al. 2003; Al-Abdulrazzak and Trombulak 2012). It is known that im-plementation of MPAs is only effective if boundaries are drawn to adequately incor-porate all important areas used by the pro-tected species, such as foraging and breed-ing areas (Peckham et al. 2007; Maxwell et al. 2011). However, significant gaps often remain in the design and functioning of MPAs (Pullin et al. 2004; Sale et al. 2005), many of which rarely present an adequate size to preserve a representative sample of regional biodiversity, or to provide ade-

instance, about one third of the mangroves, seagrass beds and wetlands worldwide have been lost in recent decades as a result of deforestation and urbanisation (Lewis et al. 2011; Penha-Lopes et al. 2011). Coral reefs, the most biodiverse ecosystems in the ocean, are estimated to harbour around one third of all described marine species (Reaka-Kudla 2001). These reefs are highly threatened and their loss would mean the extinction of much of the world’s total ma-rine biodiversity (Veron et al. 2009).

Perhaps owing to its wide geographic range and habitat connectivity, there is a wide-spread perception that extinction in the oceans is unlikely (Roberts and Hawkins 1999; Hendriks et al. 2006). Far from this perception, species and marine habitats are increasingly endangered; in fact, in the last decade more than 100 extinctions of marine populations at the local, regional and global scale have been compiled (Dulvy et al. 2003). Although research on biodiversity conserva-tion has increased in recent decades, these efforts are dominated by studies on terres-trial ecosystems, while marine environments remain largely unexplored (Carr et al. 2003; Leslie et al. 2003; Hendriks et al. 2006; Scott et al. 2012). These differences in re-search turn into differences in conservation measures carried out in each environment. Currently less than 1% of the world’s seas are under any form of protected area des-ignation, while protected areas on land cov-er 11% of the earth’s land surface (Naugh-ton-Treves et al. 2005; Toropova et al. 2010). Paradoxically, the fact that oceans are the patrimony of all nations creates a legislation gap and thereby causes a major obstacle to boost the percentage of protected surfaces in the oceans (Norse 2010).


made important contributions to their man-agement and conservation in terms of habitat range and designation of efficient protected areas for them. The effectiveness of extant MPAs in protecting important habitats for marine turtles has been assessed in several re-cent studies, some of which have underscored the relative effectiveness of these areas since core use areas of many populations largely exceed their boundaries (Maxwell et al. 2011; Nel et al. 2013).

1. 5. MARINE TURTLESModern marine turtles belong to an ancient group of reptiles inhabiting the Earth for over 110 million years, since the Cretaceous (Hirayama 1998). From the Cretaceous, four families of marine turtles diverged: Protoste-gidae; Toxochelyidae; Dermochelyidae; and Cheloniidae, but only the latter two have sur-vived until present (Spotila 2004). Nowadays, marine turtles comprise seven extant species grouped into two families: Dermochelyidae, with the leatherback (Dermochelys coriacea) as the single extant species, and Cheloniidae, with six species: hawksbill (Eretmochelys im-bricata), Kemp’s ridley (Lepidochelys kempii), olive ridley (Lepidochelys olivacea), loggerhead (Caretta caretta), green (Chelonia mydas), and flatback (Natator depressus) turtles (Pritchard 1996). With the exception of Kemp’s rid-ley, restricted mainly to the North Atlantic and Gulf of Mexico, and the flatback turtle, endemic to the Australian continental shelf, marine turtles are circumglobally distribut-ed. They inhabit nearly all oceans, occupy-ing unique ecological niches, and exhibiting intra-specific variations in population sizes and trends, as well as reproduction and mor-phology (Wallace et al. 2011).

quate protection for species or populations with complex life histories and large area requirements (Gerber et al. 2005).

In developing countries, weak enforcement of regulations relating to protected areas is a common factor. Furthermore, not all pro-tected areas have management plans, and the national authority responsible for their pro-tection is frequently under-funded and un-der-staffed (Buitrago et al. 2008). For instance, many marine protected areas in the Caribbe-an suffer from a lack of infrastructure, insti-tutional support, enforceable regulations, and scientific information to guide management policies (Guarderas et al. 2008).

Working in the establishment of MPAs for the protection of marine vertebrate, a prio-ri knowledge of spatial habitat-use patterns helps to prioritise area-based protection strat-egies (Costello et al. 2010; Block et al. 2011). Particularly for marine turtles, we should bear in mind that these species depend as much on marine habitats as on coastal and terres-trial habitats, moving periodically from one to another. Hence, the conservation of these species relies on the adequate protection of these two types of environments. It is obvi-ous that you cannot protect something if you are unaware of its location; this is especially true for the establishment of protected areas for highly migratory species such as marine turtles, whose foraging grounds and suitable breeding habitat are separated by hundreds or thousands of kilometres. In the last de-cade, our understanding of movements and migrations, the use of preferred areas, and spatio-temporal patterns of habitat use of marine turtles have been widened by the use of satellite telemetry (Godley et al. 2008; Cos-ta et al. 2012). Satellite tracking of marine turtles, as well as in other marine wildlife, has


1997). From six weeks to two months later (depending on the species), hatchlings make their way to the surface of the sand and head to the water. Hatchlings are transported by ocean currents to oceanic habitats, where they live in flotsam, such as Sargassum mats and have an omnivorous diet. Carr (1987) hypothesized that hatchlings spend their first years in oceanic habitats presumably feeding primarily on sea jellies and salps. This period of time is often referred to as the “lost years”; but a recent study based on satellite telemetry methods has shed light on this period (Mansfield et al. 2014). After this oceanic period, they return to coastal waters where they forage and continue to mature. Once adult males and females ac-quire sufficient resources, they migrate to breeding areas to mate. The time it takes to reach sexual maturity (when they are able to reproduce) varies among species, but ranges between approximately 10-30 years. The distance between feeding and breeding areas can be hundreds, to tens of thousands kilometres, with turtles performing seasonal migrations moving across large expanses of the marine environment.

1. 5. 1. Life cycle Marine turtles depend on both marine and terrestrial habitats for their growth and development, from high energy beaches to benthic reefs, and the open waters of the seas. The seven species have similar life cycles (Figure 1.1) with variations in the duration of phases (Miller 1997). Nesting females are philopatric to natal regions with sexually mature animals returning to their natal beaches to breed and nest, and both males and females can be philopatric to breeding areas adjacent to a nesting beach (FitzSimmons et al. 1997; Velez-Zuazo et al. 2008). However, a certain degree of variations in philopatry among populations and species has been described.

In general, female marine turtles typical-ly nest more than once per reproductive season. They do not nest every year and their nesting behaviour is highly stereotypic (Meylan and Meylan 1999). Only females will come ashore to dig a hole into which they deposit between 50-200 soft-shelled eggs, depending on the species (Miller

Figure 1.1. Illus-tration of a gener-alized life cycle of marine turtles. (Re-drawn from Miller 1999, In: The Biol-ogy of Sea Turtles, Kennish MJ and Lutz PL eds, CRC Press, Boca Ratón, New York).


1. 5. 2. Roles of marine turtles in ecosystems

Marine turtles are recognised as a “keystone species” because of their ecological impact on their ecosystem structure and function (Bjorndal and Bolten 2003). Marine turtles play fundamental ecological roles in ocean ecosystems by maintaining healthy seagrass beds and coral reefs, providing a key habi-tat for other marine life and facilitating nu-trient cycling from water to land (Bjorndal 2003; Bjorndal and Jackson 2003). Green turtles feed primarily on seagrass (Thayer et al. 1982) and hawksbill turtles have a di-etary preference for marine sponges. This can have a positive indirect effect on cor-als by grazing on coral competitors, and it can affect overall reef benthic biodiversity (León and Bjorndal 2002). The leatherback turtle is a globally significant consumer of jellyfish, playing an important ecological role as a top jellyfish predator (Houghton et al. 2006). Turtles host parasites and patho-gens and are substrates for many species of epibionts (Sazima and Grossman 2006; Frick 2013). Moreover, they can improve their nesting beaches by supplying a con-centrated source of high-quality nutri-ents from distant and dispersed foraging grounds (Meylan et al. 1995; Bouchard and Bjorndal 2000).

1. 5. 3. Main threats to marine turtlesMany marine turtle populations have been subject to high levels of harvesting and oth-er indirect threats and all species for which data are available are now of conservation concern (IUCN 2013). Threats vary across regions, but general categories include fisheries bycatch (i.e. incidental capture by

marine fisheries operations targeting other species), exploitation of eggs, meat or other turtle products, coastal development, pol-lution and pathogens, and climate change (Wallace et al. 2011). These threats occur at all stages of their life cycle. Anthropo-genic threats such as the slaughter of turtles for their meat and egg take are the main threats at nesting beaches. Other sources of egg mortality are bacterial and fungal in-fections (Patiño-Martínez et al. 2011), egg predation by ghost crabs, ants and other animal species (Blamires et al. 2003; Caut et al. 2006) or non-natural predators such as feral dogs, pigs, raccoons or mongooses (Ordoñez et al. 2007; Leighton et al. 2008). Eggs and hatchlings can also be threatened by toxic chemicals, such as heavy metals and organochlorine compounds (Roe et al. 2011). Coastal development has affected nesting beaches, with many of these beach-es around the world being increasingly de-veloped for human recreation, leading to nesting habitat loss in many places. Once at sea, sharks are the primary natural pred-ator of juvenile and adult marine turtles. Currently, however, bycatch from long-line and trawling activities constitutes the main threat for these species (Wallace et al. 2013 and references therein).

In addition, global climate change could have significant negative effects on the sur-vival of many marine turtle populations. Firstly, the predicted sea level rise could be devastating to many islands, compro-mising availability of nesting beaches due to extensive coastal flooding, inundation of low-lying coastal areas and heightened coastal erosion (Fish et al. 2008). Secondly, temperature is of profound importance as an environmental factor for marine turtles,

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female (Wibbels 2003). Climate change can thus lead to a feminisation of some popu-lations, putting them at serious risk (Witt et al. 2010).

1. 5. 4. Marine turtles conservation and management

In order to protect threatened marine tur-tle nesting populations, a broad range of conservation programmes have been estab-lished around the world in which clutches and hatchlings are protected by moving eggs to enclosed hatcheries or to polysty-rene boxes (Pritchard et al. 1993; Brown et al. 2012; Maulany et al. 2012). This prac-tice has the potential risk of altering incu-bation temperatures and skewing sex ratios, often by cooling the eggs and thus increas-ing the production of males (Wibbels 2003; Maulany et al. 2012). Likewise, extreme temperature variations can negatively af-fect the hatching and emergence success of clutches (Özdemir and Türkozan 2006). Since hatcheries may have negative effects on hatching success and natural sex ratio, marine turtle eggs should be incubated in the natural nest at the laying site. It has been proposed that the relocation of eggs to a protected hatchery should be under-taken only as a last resort and only in cases where in situ protection is impossible (Mor-timer 1999). However, conservation proj-ects in areas hosting highly depleted pop-ulations under intense pressure from the illegal poaching of clutches have no other option but to incubate clutches artificially. Lack of funding is also an issue. If these conditions lead to the use of hatcheries as the only conservation strategy, it is essential to assess its effects on hatching success and the resulting sex ratios.

since it could alter the intra-annual timing of nesting as well as generate large biases in offspring sex ratio, a critical life history trait (Hawkes et al. 2009; Hulin et al. 2009).

Any threat involving temperature changes is crucial for the survival of marine turtles since, as many reptile species, they exhibit temperature-dependent sex determination (TSD) where primary sex ratio is influenced by the temperature experienced by eggs during incubation (Bull 1980). Phenotypic sex in marine turtles is determined by the temperature prevailing in approximately the middle third of the incubation period (Yntema and Mrosovsky 1982). By defini-tion, the pivotal temperature is the tem-perature at which both sexes are produced in equal proportions (sex ratio= 1:1). The transitional range of temperatures (TRT) is the range of constant temperatures that yields both sexes in variable proportions (Mrosovsky and Pieau 1991). From this TRT, lower temperatures will produce only males and higher temperatures will pro-duce only females. These parameters are inferred from the artificial incubation of eggs at constant temperatures at the lab-oratory (Mrosovsky et al. 2009). Where pivotal temperatures are known, incuba-tion temperatures can be used to predict hatchling sex ratios (Mrosovsky et al. 2009). Likewise, differences in nest temperatures manifest themselves as extended or short-ened incubation durations (ID), for cooler and warmer conditions respectively, and thus ID can be used to some extent to infer sex hatchling ratios (Mrosovsky et al. 1999). In nesting beaches around the world there is a predominance of beaches that produce female-biased hatchling sex ratios, and some of these biases are greater than 90%


al activities, increased river sediment load-ing, the introduction of alien species and climate change are identified to be among the major sources of anthropogenic pres-sure on Caribbean marine life (Burke and Maidens 2004).

Six species of marine turtles occur in the Caribbean: loggerhead turtles, green tur-tles, hawksbill turtles, Kemp’s ridley turtles, olive ridley turtles, and leatherback turtles. Each of these species is classified by the World Conservation Union (IUCN 2013) as either critically endangered (hawksbill and Kemp’s ridley), endangered (green and loggerhead turtles) or vulnerable (leather-back and olive ridley). The use of a global extinction risk assessment framework repre-sented by the Red List for the assessment of that species status at the regional scale has been challenged by certain experts (Seminoff and Shanker 2008; Wallace et al. 2011). According to these specialists, this type of assessment will be more useful for conservationists and on-the-ground re-source managers. Regardless of the frame-work used, it’s a given that populations of endangered Caribbean marine turtles are far more depleted than previously thought because current conservation assessments do not reflect historical nesting data. In the past, large nesting populations were found on beaches throughout the wider Caribbe-an (McClenachan et al. 2006) but, nowa-days large nesting colonies are rare (Figure 1.2). Nesting grounds receiving more than 1,000 crawls per year range from 0.4% (hawksbill) to 7.0% (Kemp’s ridley) of all known species-specific sites. For any spe-cies, roughly half of all known nesting sites support fewer than 25 crawls (fewer than 10 reproductively active females) per year

1. 6. MARINE TURTLES IN THE CARIBBEAN REGION

The Wider Caribbean Region consists of twenty-eight sovereign nations and com-prises nine ecoregions: Western Caribbean, Southwestern Caribbean, Eastern Carib-bean, South Caribbean, Greater Antilles, Bermuda, Bahamian, Southern Gulf of Mexico, and Florida (Spalding et al. 2007). This coastal area is a large marine ecosys-tem characterised by coral reefs, mangroves, and seagrass beds, but including other envi-ronments, such as sandy beaches and rocky shores (Miloslavich et al. 2010) and it is also known for its species diversity and patterns of endemism (Spalding and Kramer 2004). The Caribbean islands encompass a bio-diversity hotspot exceptionally important for global biodiversity conservation, due to high levels of species endemism (Roberts et al. 2002). In terms of endemism at the ge-nus level, it ranks third among the world’s 34 biodiversity hotspots, with 205 plants and 65 vertebrate genera endemic to the islands (Smith et al. 2004).

There is a complex mix of interacting so-cio-economic, political, cultural and envi-ronmental factors that are driving environ-mental change and threatening biodiversity in the Caribbean (Anadón-Irizarry et al. 2012). The region has suffered from high levels of habitat loss since the arrival of Eu-ropeans in the 1490s. This destruction has reduced the hotspot’s original estimated 229,549 km2 of natural vegetation to just 22,955 km2 (or just 10%). Nowadays, rising population densities and associated rapid, unchecked coastal development, increasing fishing pressure, agricultural and industri-

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(Dow et al. 2007). Calculations based on historical export data show that modern populations of green and hawksbill turtles

are 0.33% and 0.27% of their historical numbers, respectively (McClenachan et al. 2006).

Figure 1.2. Map of nesting habitats (red dots) and legal status of marine turtles within Exclusive Economic Zones (EEZ) for the Caribbean countries (modified from Dow et al. 2007). Blue areas include some protection but allow traditional uses.

tortoiseshell products around the world has deeply influenced the survival status of the species (Meylan 1999). In the Caribbean, large-scale commercial harvesting and trade began in the 1950s and 1960s, when international markets for shells from hawksbill and other turtle species expanded (Groombridge and Luxmoore 1989). Despite the restricted international regulations to protect them (all of these species are listed on CITES Appendix I, which prohibits international commercial trade), there is still a large amount of trade in hawksbills products (Bräutigham and Eckert 2006). The Caribbean stands out as a region of exceptionally high levels

Marine turtles in the Caribbean have been subject to exploitation for many centuries, which has resulted in the large-scale reductions in population numbers stated above (Parsons 1962; Eckert 1995; Jackson 1997; Meylan 1999; Bell et al. 2006). Turtle eggs and most turtle body parts (meat, shell and skin) have been valued not only because they provide basic sustenance but also because they have been used to create jewellery for their trade (Fleming 2001; Bräutigam and Eckert 2006). This is particularly dramatic for the hawksbill turtle, sought after for its scutes which have been long valued as raw material for artisans. The intensity of the demand for

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in regional marine turtle conservation and should be considered as potential priorities for management and conservation (Dow and Eckert 2011). In agreement with Mc-Clenachan and colleagues (2006), it is better to extend protection across as many beaches as possible so as not to rely on the protection of few exceptional nesting beaches.

1. 7. MARINE TURTLE CONSERVATION: THE CASE OF THE DOMINICAN REPUBLIC

The Dominican Republic (DR) with an extension of 48,670 km2 occupies the east-ern portion of the Hispaniola Island, the second largest in size within the Greater Antilles and it is part of one of the most important biodiversity hotspots in Central America. The country supports exception-ally diverse ecosystems, but this goes hand in hand with depletion patterns as observed in wide areas devastated by deforestation and human encroachment (Myers 1988). The country includes a total of 86 protect-ed areas encompassed in six categories with different levels of protection, including one Biosphere Reserve (Jaragua-Bahoruco-En-riquillo) and 19 national parks managed by the National Parks Directorate. Although none of them were created specifically for protection of marine turtles, some provide suitable nesting and foraging habitats.

Four marine turtle species have been re-ported in the DR’s coastal areas: hawksbill turtle, green turtle, loggerhead turtle and leatherback turtle (Ottenwalder 1981).

of tourism development which has led to irreversible environmental degradation (Davenport and Davenport 2006). This development, combined with increasing pollution, sand extraction, sewage and litter have all been detrimental to turtle nesting habitats in the region (Bell et al. 2007).

Based on a recent study carried out to estab-lish conservation priority in different marine turtle Regional Management Units (RMUs) (Wallace et al. 2011), “Low risk-High threats” is currently the most prevalent con-servation priority category for marine turtles in the Wider Caribbean region. This cate-gory highlights RMUs generally exhibiting large, stable or increasing abundance pop-ulations, yet highly under threat. If threats are not abated, these populations could decline in the future, thus warranting inter-vention before significant population-lev-el impacts can manifest. The leatherback Northwest Atlantic RMU is currently cate-gorised as “Low risk-Low threats” meaning large populations that, in many cases, are well-monitored. This classification supports recent results on marine turtle population status in the region, which quoted that some populations are stable or increasing (Leath-erback: Dutton et al. 2005; McGowan et al. 2008; hawksbill: Richardson et al. 2006; Kamel and Delcroix 2009; green: Troëng and Rankin 2005). Although recent conser-vation efforts have resulted in large popula-tion increases at several nesting sites, there are still many gaps on the conservation sta-tus of many small, widely dispersed sites lacking intensive population monitoring; many of these unprotected rookeries in this basin are seriously threatened (Lagueux and Campbell 2005; Dow et al. 2007). These populations could play an important role

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ternesting period (Whiting et al. 2006). In foraging grounds juveniles converge from many different nesting aggregations, each consisting of genetically distinct subpopu-lations (Bass et al. 1996; Velez-Zuazo et al. 2008). Therefore, these are areas of great interest for the conservation of the species.

Leatherback turtleThe leatherback turtle is the largest of all living turtles; adults can reach more than 2 meters in total length and often exceed 500 kg (Figure 1.3b). The lyre-shaped cara-pace has seven longitudinal ridges, or keels, and black colouration with white spots. The leatherback turtle is an air-breathing diving animal capable of maintaining ac-tivity during prolonged dives (with a max-imum of 1280 m depth recorded) fuelled by its oxygen stores (López-Mendilaharsu et al. 2009). The leatherback turtle has a worldwide distribution, spreading from tropical to sub-polar oceans and nesting on tropical (rarely subtropical) beaches. De-spite its extensive range, distribution is far from uniform and large nesting colonies are rare. In the Western Atlantic, nesting occurs as far north as Assateague Island National Seashore, Maryland, USA (38ºN) (Rabon et al. 2003), and as far south as Tor-res, Rio Grande do Sul, Brazil (29ºS) (Soto et al. 1997). The largest nesting colonies are located in French Guiana-Suriname (over 40% of the world leatherback popu-lation nests, Hilterman and Goverse 2007). Leatherbacks spend the first part of their lives in tropical waters, and once they ex-ceed 100 cm CCL they are considered to be sub-adults and are able to move into cooler waters that have been considered the pri-mary habitat for the species. Leatherback

Hawksbill turtleThe hawksbill turtle is medium size, with a straight carapace length between 60 and 90 cm in the adult stage (Figure 1.3a). The shell is elongated and has overlapping scutes and a serrated edge. This is the most tropical of all marine turtles, with nesting beaches and feeding grounds distributed in tropical and subtropical areas of the At-lantic, Pacific and Indian Oceans between 30º N and 30º S (Baillie and Groombridge 1996). In the western Atlantic Ocean, Gulf of Mexico, and the Caribbean Sea, hawksbills are found from the southern U.S. southward along the Central Ameri-can coast to Brazil and throughout the Ba-hamas and the Greater and Lesser Antilles (Meylan and Redlow 2006). Hawksbill tur-tles forage in a variety of coral and sponge reefs, reef walls, and other hard-bottom habitats throughout the tropics. Sponges are the main component of their diet but they can also include substantial quantities of non-sponge invertebrates that may be present in their feeding habitat (León and Bjordnal 2002). Through their selective foraging behaviour on sponges, hawksbills play a crucial role in the conservation of reef communities favouring reef succession and diversity. Tagging and satellite tracking studies conducted throughout the Caribbe-an have demonstrated that hawksbills carry out reproductive migrations through the territorial waters of multiple jurisdictions, crossing oceanic areas and moving through other neritic areas to their final foraging grounds (van Dam et al. 2008). Hawksbills nest on both insular and mainland sandy beaches, often in areas with at least some vegetation (Kamel and Mrosovsky 2005). Nesting females typically remain within the vicinity of the nesting beach in the in-


cies fills unique ecological roles in seagrass ecosystems by reducing the flux of organic matter and nutrients to sediments. Their removal from this ecosystem would lead to deposition of plant detritus, increase the oxygen demand of sediments, and promote hypoxia (Jackson et al. 2001). The primary green turtle nesting rookeries (i.e. sites with ≥ 500 nesting females per year) are located at Raine Island, Australia (Limpus 2007) and Ascension Island (Mortimer and Carr 1987). In the Caribbean region, the largest remaining green turtle rookeries are locat-ed at Tortuguero, Costa Rica and Aves Is-land, Venezuela (Seminoff 2002).

Loggerhead turtleThe loggerhead turtle is a medium-sized turtle; mature females have a mean straight carapace length of 87 to 105 cm; and a mean weight near 115 kg (Spotila 2004) (Figure 1.3d). They have a characteristic large head with a very strong neck and powerful jaws. The carapace is slightly heart-shaped and reddish-brown in adults and sub-adults, while the plastron (ventral carapace) is gen-erally a pale yellowish colour. Loggerhead turtles are generalists in feeding throughout their lives. In oceanic habitats hatchlings feed on crabs, molluscs, jellyfish and Sar-gassum (Bjorndal 1997; Bolten 2003). As ju-veniles, they enter in a benthic feeding stage into estuaries, lagoons and other coastal re-gions where they consume hard-shelled in-vertebrates (Dodd 1988; Hopkins-Murphy et al. 2003). However, a dichotomy in feed-ing strategy has been observed for the spe-cies, with juveniles using different habitats for development, thus resulting in adults of the same rookery foraging in coastal waters and the others into oceanic areas (Hatase et

turtles perform long distance migrations (1000s of km) that may span as long as 2-3 years, from nesting beaches to high latitude foraging grounds (Eckert et al. 2006; Hays et al. 2006). In the foraging grounds they feed primarily on soft-bodied animals like jellyfish and pelagic tunicates (Heaslip et al. 2012). In the Atlantic, satellite tracking data has revealed site fidelity for foraging grounds across the North Atlantic (includ-ing Canada, the northeastern US and West-ern Europe), and West Africa (e.g., Ferraroli et al. 2004; Hays et al. 2004; 2006, James et al. 2005; Doyle et al. 2008).

Green turtleThe green turtle (Figure 1.3c) is a circum-global species occurring throughout tropi-cal and, to a lesser extent, subtropical wa-ters. They are believed to inhabit coastal waters of over 140 countries (Groombridge and Luxmoore 1989). The typical adult has a carapace length reaching 120 cm and can weigh as much as 230 kg in the Atlantic, making this species the largest of the hard-shelled marine turtles (Pritchard and Mor-timer 1999). The carapace is broadly oval and the margin is sometimes scalloped but not serrated. Dorsal carapace colour varies among populations, from black to grey-brown spotted patterns. Ventral colour is lighter. The green turtle is the only herbivo-rous marine turtle foraging primarily on sea grasses in most of its range in large juvenile and adult stages (Hirth 1971), with algae making up the bulk of the diet where sea-grasses are lacking (Bjorndal 1985). Green turtles exhibit particularly slow growth rates and the age to maturity for the spe-cies appears to be the longest of any ma-rine turtle species (Hirth 1997). This spe-


al. 2002; Hawkes et al. 2006; Watanabe et al. 2011). Loggerheads are a circumglobal species, occurring throughout the temper-ate and tropical regions of the Atlantic, Pacific, and Indian Oceans, nesting most abundantly in subtropical and temperate areas, and occasionally in the tropics. The world’s largest populations of loggerhead turtle nesting colonies are located in Florida

(Spotila 2004) and Masirah Island, Oman, in the Indian Ocean (Rees et al. 2010). In the Wider Caribbean, important nesting grounds for this species are located along the southeastern coast of the USA, (mainly in the state of Florida), Brazil, the Yucatan Peninsula in México, Cuba and the Colom-bian coast (Dow et al. 2007).

Figure 1.3. Marine turtle species reported in the literature in coastal areas of DR. a) Eretmochelys imbricata, b) Dermochelys coriacea, c) Chelonia mydas, d) Caretta caretta. (Photos courtesy of: a, b, c: J. A. Álvarez and d: J. Tomás.

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Historically, nesting turtles were abun-dant in the Dominican Republic (Parsons 1962). Before the arrival of the Europeans to Hispaniola, marine turtles were a prom-inent element in the existing native cultures (Mota and León 2003). Since the arrival of Columbus, there have been many referenc-es about the great presence of marine tur-tles as well as how appreciated their meat and eggs were (Rodríguez Demorizi 1942; Parsons 1962). The Dominican Republic has a long history of harvesting and com-merce in marine turtles, which have con-stituted an important resource for coastal communities (Ottenwalder 1981; Fleming 2001; Reuter and Allan 2006). The hawks-bill’s meat, eggs, penis, blood, shell and heart have a known history of use in the Dominican Republic (Figure 1.4a to 1.4e). Green and leatherback’s meat, eggs and fat are also consumed. Harvesting practices to supply the Dominican shell industry have represented a serious threat to the species’ regional survival outlook. Widespread sales of tortoiseshell items in Santo Domingo and coastal localities have been well docu-mented in shops catering to tourists (Stam and Stam 1992; Dominguez and Villalba 1994; Feliz et al. 2010; Figure 1.4b and 1.4c). Habitat loss due to coastal develop-ment constitutes a second important threat to marine turtle nesting habitats (Otten-walder 1996) (Figure 1.5).

Figure 1.4. Human interaction with marine turtles in Dominican Republic. a) The carapace of a nesting hawksbill captured by humans in Saona Island, b and c) Different types of accessories made from hawksbill shell for sale in gift shops of Santo Domingo (DR), d) Skull and plastron of hawksbill turtle captured by humans in Saona Island, e) Juvenile hawksbill turtle entangled in a fishing net. (Photos courtesy of: a, b and c: Y.M. León, d: J. Tomás and e: S. Aucoin).

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Marine turtles are now legally protect-ed and their trade banned in the country by laws dating back to 1966, and recently confirmed through the 2005 Fisheries Law CODOPESCA (Dominican Council of Fisheries and Agriculture). However, the Dominican Republic and other Caribbean islands share a poorly established environ-mental policy and an institutional weakness (Torres et al. 2000). Therefore, despite the existence of laws and protected areas in the country, “poor supervision” and committed offenses to environmental laws are frequent. Tourism expansion on the coast has led to an appropriation of coastal pieces of land, including territories of coastal and marine protected areas (Heredia 2003). Moreover, tourism constitute a serious threat caus-ing environmental degradation due to sol-id waste, uncontrolled tourist influx to the beaches or illegal construction on coastal areas (Figures 1.5 and 1.6).

Figure 1.5. a) and b). Debris covering a leatherback nesting beaches at the east of Jaragua National Park (SW DR). (Photos courtesy of: J. Tomás).

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Previous knowledge about the conservation status of marine turtle nesting populations in the Dominican Republic relies on the studies carried out by Ottenwalder (1981) and Ross and Ottenwalder (1983) based on surveys along the coast and its small islands. They identified the areas of special interest for nesting leatherback, green and hawks-bill turtles. Moreover, they provided the first data on abundance, reproductive biology and conservation status of these species in the DR.

No exhaustive studies on marine turtle nest-ing population have been carried out during the last three decades; however, there have been numerous reports condemning the alarming level of marine turtle harvesting and commerce in the country (Stam and Stam 1992; Dominguez and Villalba 1994). The lack of updated data on the situation of these species and the threats reported have made it necessary to carry out an ex-haustive study in the area not only for the status of Dominican turtles themselves but also for the repercussion that the lack of knowledge could have upon the conserva-tion of marine turtles at the regional level due to the links between populations, the different habitats used by nesting females and their ecological role in Caribbean ecosystems (Bräutigham and Eckert 2006; Dow and Eckert 2011).

Figure 1.6. a) Mass tourism at Bayahibe village (Del Este National Park, SE DR), b) Off-road vehicle at the protected marine turtles nesting beach of Bahía de las Águilas (JNP), and c) Coastal development at Del Este National. (Photos courtesy of: J. Tomás, Y.M. León and O. Revuelta).

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Marie Curie PEOPLE-2009-RG FP7) and the General Foundation of the University of Valencia. All these projects have been a coordinated joint effort between the Uni-versity of Valencia, the Dominican NGO Grupo Jaragua, the University of Exeter (United Kingdom), the University of San-to Domingo (through the Research Center for Marine Biology - CIBIMA-and School of Biology), the Technological Institute of Santo Domingo (INTEC), and the Arau-caria Project of the Spanish Agency for International Cooperation (AECI), with continuous support from the Dominican Ministry of Environment and Protected Areas. During these five years, all these projects have covered research on repro-duction biology and threats (in the present thesis), public awareness and training, but also other studies such as identification of feeding areas of nesting females through satellite telemetry (Hawkes et al. 2012) and populations genetics (Carreras et al. 2013).

1. 8. THE PROJECTThe present study has been carried out within the framework of a five-year project on the study of marine turtle nesting pop-ulations in the Dominican Republic. The project was initiated as a pilot study to as-sess marine turtle populations and the main threats on beaches and coastal waters of the Jaragua National Park. The first studies had the support of a project (PCI-A/2991/05) awarded to the University of Valencia (Spain) by resolution of the Spanish Agen-cy for International Cooperation (AECI) published on 5 January, 2006. During this year, training of local people and public awareness were also included in an attempt to ensure long-term data gathering and conservation legacy. Moreover, possibilities for the development of ecotourism were explored to generate alternative income sources to help prevent direct exploitation of these species. The project was conduct-ed by the Marine Zoology Unit (MZU) of the Cavanilles Institute of Biodiversity and Evolutionary Biology (ICBiBE) with the participation of Dominican institutions, such as the Autonomous University of San-to Domingo (UASD) and the Dominican NGO “Grupo Jaragua”, and with the col-laboration of Dominican Ministry of Envi-ronment and Protected Areas. This project was renewed in 2007, by granting the same team (PCI-A/5641/06). In the same year, the team of the MZU started the research project CGL2006-02936-BOS (of the Plan Nacional I+D+i, Ministry of Education and Science), that was extended until Sep-tember 2010. Additional funding has been obtained from the European Union (proj-ect IEOST, Marie Curie Intra European Fellowship FP6, and Project RESET-ECO,

2. CHAPTER II: AIM AND OBJECTIVES©

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ChApTEr II: AIM AND OBJECTIVES

Chapter II: Aim and objectives I 67

AIMThe approach advocated in this thesis is based on the fact that there is a gap of in-formation regarding the conservation status of marine turtle nesting populations in the Dominican Republic. It is recognized that marine turtle conservation requires ample knowledge of different aspects such as re-productive biology or habitat-use patterns. Developing tools and measures for their protection, as well as evaluating the results of such measures are also essential. This thesis is focused on identifying the main marine turtle nesting rookeries in the Do-minican Republic, describing the current spatio-temporal patterns of nesting, and assessing the likely impact of the current threats to these nesting stocks, spanning a 5-year study period (2006 - 2010).

OBJECTIVESThe specific objectives of this study are:

1. To identify the main breeding colonies of marine turtles in the Dominican Republic, describing the spatio-temporal patterns of nesting at present and evaluating the im-pact of current threats to these breeding populations, comparing their status with previous studies in the country.

2. To evaluate the conservation programme based on artificial incubation of eggs, estab-lished for the protection of leatherback turtle clutches on the beaches of Jaragua National Park (southwestern area of the Dominican Republic), through the study of the potential effects of this artificial incubation on hatch-ing success and hatchling sex ratio.

3. As in the previous objective, to study the effect of artificial incubation on hatching success and sex ratio to evaluate this mea-sure for the conservation of the hawksbill turtle on Saona Island.

4. To study hatching success and spatial, temporal and reproductive factors affecting it to determine the reproductive value of key beaches of the Jaragua National Park for the leatherback turtle. To highlight the role of the leatherback turtle nesting on threatened beaches of the Jaragua Nation-al Park as a vehicle for the conservation of ecosystems and management in this pro-tected area.

5. To investigate the habitat-use patterns of hawksbill turtle nesting in the Domini-can Republic during their internesting and foraging periods, identifying core-use areas and comparing them with current marine protected areas in the Caribbean region.

3. CHAPTER III: GENERAL MATERIALS AND METHODS©

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ChApTEr III:GENERAL MATERIALS AND METHODS

Chapter III: General materials and methods I 71

3. 1. STUDY AREADominican Republic (DR) shares with Hai-ti the island of Hispaniola which is part of the Greater Antilles archipelago in the Caribbean region. The country is the sec-ond largest Caribbean nation (after Cuba), with 48,445 km2 and an estimated 10 mil-lion people. It has 1,576 km of coastline, including islands, islets and cays; 526 km (33%) on the north coast, 374 km (24%) on the east coast and 675 km (43%) on the south, with a surface of 9,484 km2 of continental shelf. The country has trop-ical climate, with a strong maritime influ-ence controlling general weather patterns, apart from the predominant influence of the trade winds. The average temperature is 25 °C (average at coast level: 28 °C) and precipitation ranges from 350 mm to 2.500 mm. Hurricane season in the country runs from June to November. The National Sys-tem of Protected Areas has 86 units encom-passed in six categories: Strictly Protected Areas (n= 8), National Parks (n= 19), Nat-ural Monument (n= 19), Habitat/Species Management Area (n= 13), Strict Nature Reserve (n= 15) and Protected Landscape/Seascape (n= 12).

Jaragua National ParkJaragua National Park (JNP) in the south-western part of the DR covers 1,374 km2, of which 905 km2 are marine area (Fig-ure 3.1). The Park includes the islands of Beata and Alto Velo, as well as Los Frailes and Piedra Negra cays. The JNP was es-tablished on August 11, 1983 by Presiden-tial Decree and, since 2002, it is one of the core areas of the Jaragua-Bahoruco-En-riquillo Biosphere Reserve. The Park in-

cludes numerous ecosystems, ranging from natural forests, beaches, rocky shores, wet-lands, seagrasses and coral reefs, with very high levels of flora and fauna endemism. In general, the vegetation is characterized by a great variety of plants adapted to high solar radiation and low precipitation. Its marine ecosystems hosts one the most ex-tensive and best preserved sea grass beds in the southern coast, which supports many threatened species.

The Park comprises two groups of beach-es about 50 km apart that provide nesting areas for leatherback (Dermochelys coriacea) hawksbill (Eretmochelys imbricata) and green (Chelonia mydas) marine turtles. At the north-eastern side of the Park, on the narrow area of land between Oviedo Lagoon and the Caribbean Sea, are located the beaches of San Luis (11 km length), Mosquea (3.3 km) and Inglesa (1.2 km) (Figure 3.1). These beaches have coarse, dark sand, a steep slope and strong wave action and sea cur-rent influence which contributes to the ac-cumulation of large amounts of plastic and other debris on their shores. The western-most beaches, Bahía de las Águilas (4.4 km) and La Cueva (2.5 km), have fine-grained, coralline, white sand formed by the coral reefs that are nearshore (Figure 3.1). The Park includes other beaches that probably were used for marine turtle nesting in the past, but where currently nesting is scarce.

There are six human settlements with a to-tal population of ca. 5000 located around the boundaries of the Park. Moreover, the Park receives visits of people from the close town of Pedernales and itinerant fishermen either from other parts of DR and from the nearby Haiti. Although these beaches have no infrastructure development, the Park


receives 24000 visitors per year, most of them visiting the western beaches. In recent years, the tourism development of Bahía de las Águilas has been a source of debate on whether mass tourism and mining should be established to enhance the economy of the region (Wielgus et al. 2010).

Figure 3.1. Jaragua National Park limits (green line), indicating the beaches of La Cueva and Bahía de las Águilas in the west and the beaches of San Luis, Mosquea and Inglesa in the east (orange lines). The insets indicate the location of the main maps in the Caribbean. Green areas inside the Park are designated as national recreation areas.

Saona IslandSaona Island is included in the Del Este National Park (DENP, south-east DR), which was founded on September 16, 1975. With an area of 110 km2, Saona is the largest island adjacent to the DR (Fig-ure 3.2). For the most part, beach vegeta-tion is dominated by coconut (Cocos nucifera) plantations, sea purslane (Sesuvium portulac-astrum), sea rosemary (Suriana maritima), sea grape (Coccoloba uvifera), goat’s foot creeper (Ipomoea pes-caprae) and native grasses. The neritic sea adjacent to the nesting beach is composed of coral reefs and seagrass beds which provide prime habitat for hundreds of species of plants, birds, fish and other marine animals. Saona Island is also home for several endemic, threatened, or endan-gered plant and animal species. There is an uncontrolled access to the park by fisher-men from other parts of the country and expanding tourism industry that increases pressure on coral reefs and other natural resources.

There is one permanent human settlement in Saona, Mano Juan village, with a popu-lation of c. 300 inhabitants.


once a week, except for Inglesa beach where surveys were carried out every two weeks due to the distance and difficult ac-cess to that beach. From 2009 onwards, these beaches were daily surveyed by JNP’s rangers trained by the research team. On the western beaches of the Park 3-4 day-time surveys were carried out per week, but survey effort ensured the record of all nesting events every year. Other remote beaches at the JNP with low levels of nest-ing previously reported were also occa-sionally visited to confirm reports of nests from reliable informants. In Saona Island, beaches were patrolled at least once per week throughout the year, but in 2008 sur-vey effort was increased to 3-4 surveys per week during the period of peak nesting activity (June to November, see chapter 4).

Figure 3.2. Saona Island, indicating Mano Juan village and the four surveyed areas (black lines) of (1) El Toro, (2) Mano Juan, (3) Canto la playa and (4) Faro Punta Cana. The insets indicate the location of the main maps in the Caribbean.

3. 2. BEACH SURVEYSIn order to appropriately design nesting beach surveys, previous knowledge on nesting activity and nesting areas (Otten-walder 1981) was revised and, interviews to rangers and local fishermen were con-ducted to determine the most important beaches to survey and the period of the nesting season.

Terrestrial surveys allow counting turtle’s tracks and species identification. Day-time surveys were undertaken on foot by researchers and trained local people on the beaches of the JNP and Saona Island during five entire consecutive nesting sea-sons, 2006-2010. From 2006 to 2008, east-ern beaches of the JNP were monitored


During daytime surveys on the JNP and Saona Island, we patrolled the beaches by foot to detect all recent tracks of nesting females (Figure 3.4a and 3.4b). We record-ed total number of emergences of nest-ing females, identified the species based upon track characteristics (Schroeder and Murphy 1999; Figure 3.4c and 3.4d) and confirmed whether each activity had suc-cessfully resulted in clutch deposition, by

Surveys in other areas of the Dominican Republic were undertaken during 2006-2010. Visits, sporadic patrols and interviews were made in 11 areas of the northern coast in 2006, 2007 and 2008, and on six beaches of the eastern and southern coasts in 2008, 2009 and 2010 (Figure 3.3). Reports about other nesting events, particularly from re-sorts or beach front developments, were also compiled.

Figure 3.3. Map of the Dominican Republic indicating places where surveys and interviews have been undertaken. a-d: beaches visited with interviews performed and sporadic nesting recorded, e: Del Este National Park with Saona Island, f: Jaragua National Park, g: Del Muerto beach, also surveyed sporadically during the study period (see text for details). Black pushpins indicate other records of low level of sea turtle nesting activity communicated by others and/or occasionally visited by members of the project team and collaborators.


Figure 3.4. Different images of the surveys carried out during the study period (2006-2010). a) nesting beaches at Saona Island, b) daytime surveys on Bahía de las Águilas (western JNP), c) data collection of a leatherback turtle track at Bahía de las Águilas beach, d) hawksbill turtle track on Canto de la playa beach (Saona Island). (Photos courtesy of: Y.M. León, O. Revuelta, J. Tomás and A. Mason).

digging to find the clutch. Nest location was recorded using a global positioning system (GPS) and at-site features, and turtle tracks were marked to avoid duplicate counts (Fig-ure 3.5a). Signs of eggs taken by people were recorded (presence of probing sticks, evidence of digging and broken eggshells, and/or human footprints around the nest-ing site; Troëng et al. 2004) (Figure 3.5b and 3.5c). Remaining clutches were either camouflaged and incubated under natural conditions in the beach or, when clutches were in high risk of subsequent take (be-cause of proximity to a human settlement or presence of people in the area), they were transferred to protected places for egg incubation in boxes.

Because of funding limitations, night sur-veys were only carried out on Saona Island and the western beaches of the JNP. During night surveys, when finding a female turtle nesting the protocol was to leave someone from the survey team with the animal un-til she finished the eggs laying, while the rest of the team continued with the survey. After oviposition, turtles were measured, tagged and examined for possible injuries, epibionts or other abnormalities (see next section).

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3. 3. DATA COLLECTION AND CLUTCH RELOCATION

For every nest, we collected data on date and time of laying, GPS coordinates, nest location across the beach (distance to high tide line), nest location along the beach (sector) and beach zone (open sand, vegeta-tion border, and within vegetation). A flex-ible tape was used to obtain curved mea-sures of the turtles; curved carapace length (CCL) and maximum curved carapace width (CCWmax). Hawksbill and green turtle carapaces were measured from the anterior point at midline (nuchal scute) to the posterior notch at midline between the supracaudals (Figure 3.6a); leatherbacks were measured from the anterior edge of the carapace at the midline to the posterior tip of the caudal peduncle. Carapace width was measured at the widest point for the three species (Figure 3.6b). Hawksbill and green turtle females were tagged with small metal Inconel tags on the trailing edge of both fore-flippers (Figure 3.6c). Leather-back females were tagged with large metal Inconel tags (National Brand and Tag Co., Newport, USA) fitted between the tail and rear flippers (Figure 3.6d).Figure 3.5. a) Leatherback nest surrounded by debris

on Mosquea beach (eastern JNP), b) predated hawksbill clutch in Saona Island, c) rests of eggshells from a predated clutch. (Photos courtesy of: P. Feliz and Y.M. León).

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Once the turtle left the beach we proceeded to the management of the clutch. The fate of the clutches was decided according to the risk of flooding, presence of natural/feral predators or passers-by, and its proximity to fishing camps and villages. If no threats were found, clutches were left for incubation in the beach under natural conditions (in situ). In those cases with clear risk of clutch preda-tion, the nest was excavated and eggs trans-ferred to polyethylene exterior boxes with polyurethane foam filling (dimensions 30 cm wide x 50 cm long x 32 cm deep) for their incubation at the park rangers’ huts (Figure 3.7a and 3.7b). Two to three centimetres of nest sand was put at the bottom and sides of the box to prevent contact of eggs with walls. Clutches were carefully excavated by hand and eggs extracted avoiding their rota-tion, as has been recommended to improve the success of delayed relocations (Abella et al. 2007; Figure 3.7b). Once in the box, the top eggs were also covered with two to three centimetres of sand proceeding from the original nest chamber.

For each clutch the following data were col-lected (Figure 3.7c to 3.7e):

- Clutch size (number of yolked eggs): the number of eggs laid into the nest, excluding yolkless eggs.

- Number of yolkless eggs: small eggs (around half the diameter of yolked eggs) containing mostly albumen encapsulated by a shell.

- Egg size: Ten eggs were chosen at ran-dom from each clutch. Each egg was cleaned of adhering sand and the great-est diameter was measured to the nearest 0.1 cm with a caliper.

Figure 3.6. Measuring and tagging marine turtles. a) measuring CCL of a nesting hawksbill turtle in Bahía de las Águilas, b) measuring CCW of a nesting leatherback turtle at La Cueva beach, c) detail of tagging a nesting hawksbill turtle, and d) tagging a nesting leatherback turtle between the tail and right rear flipper at western JNP. (Photos courtesy of: Y.M. León, O. Revuelta, J. Tomás and Grupo Jaragua).

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- Egg weight: Eggs were weighed using an electronic balance read to a minimum accuracy of 0.1 g.

- Nest bottom depth: Depth from beach surface to bottom of egg chamber (cm).

Each box was labelled with a code indicat-ing beach name, laying date and number of eggs incubated. The boxes were stored in Park rangers’ huts (Figure 3.8a and 3.8b) for their incubation and were checked dai-ly throughout the incubation duration and lids opened for two to three hours a day to allow air circulation.

Figure 3.7. Images of egg collection. a) and b) Counting and transferring hawksbill eggs to a polyethylene box in Saona Island, c) Yolked and yolkless eggs in a nest chamber of a nesting leatherback turtle, d) measuring leatherback turtle eggs, e) weighing hawksbill turtle eggs. (Photos courtesy of: Y.M. León, O. Revuelta and J. Tomás).

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observed out of the nest rangers marked them. Clutches were excavated for hatching success study no earlier than 48 hours after track detection.

During the study of both in situ and artifi-cially incubated clutches the next categories were considered and quantified as follows (Figure 3.9):

- Live hatchlings: hatchlings found alive among shells.

- Dead hatchlings: hatchlings found dead out of the shell.

- Yolkless eggs: number of eggs without yolk.

- Egg shells: empty open shells remain-ing after hatchlings emerge from the nest. We counted as one shell those that make up more than 50% of the egg size.

- Unhatched eggs: Eggs with sign of embryo development. These eggs were opened and classified as:

Early stage embryonic death: eggs that had par-tial calcification of the shell or evidence of an early stage dead embryo (Bell et al. 2003).

Late stage embryonic death: eggs that contained dead embryos at late stage of development.

Figure 3.8. Rangers huts where artificial incubation was carried out. a) Saona Island and b) Bahía de las Águilas. (Photos courtesy of: Y.M. León and J. Tomás).

3. 4. HATCHING SUCCESS AND EMERGENCE SUCCESS

Once artificially incubated eggs hatched, clutches were carefully excavated no earlier than 48 hours after the last sign of hatch-ling emergency. Hatchlings were released to the sea on their origin beaches (when pos-sible) when they entered into frenzy. Before they were released, random samples of 20 hatchlings from each clutch were measured (straight carapace length, SCL) to the near-est 0.1 cm with a calliper, and weighed to the nearest 0.1 g with an electronic scale. Clutches left incubating on the beach (in situ) were monitored during the incubation duration and, when hatchling tracks were

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Figure 3.9. Clutch study. a) An exposed leatherback hatched clutch contents: 1. Yolkless eggs, 2. Yolked eggs, 3. Hatched eggshells. From b to e) Excavating and collecting data of leatherback and hawksbill clutches at study sites. (Photos courtesy of: a, b: J. Tomás, c: Héctor González, d and e: O. Revuelta).

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For the study, we defined hatching success and emergence success as follows, according to the literature (Miller 1999):

Hatching success (HS) is the number of hatchlings that hatch out of their egg shell (number of empty egg shells in the nest). HS was calculated as:

Number of egg shells

egg shells + unhatched eggsHS = x 100((

Emergence success (ES) is the number of hatchlings that reach the beach surface. ES was calculated as:

Number of egg shells - (live hatchlings + dead hatchlings )

egg shells + unhatched eggsES = x 100((

Hatchlings were released to the sea on their origin beaches (Figure 3.10).


start with the incubation of eggs at various constant temperatures, followed by calcula-tion of pivotal incubation duration and a curve relating duration to sex ratio (Godley et al. 2001). Then this laboratory curve has to be adjusted to a curve appropriate to the field data by adding the hatch-emergence interval to the incubation durations. Final-ly, the adjusted curve is used for converting values obtained in the field on incubation durations (duration from laying to hatch-ling first emergence) to the corresponding values for clutch sex ratio estimations. Since no suitable data from laboratory are avail-able for marine turtles from DR nesting beaches, we estimated hatchlings sex ratios using the conversion curve relating incuba-tion duration to hatchling sex ratio derived from artificially incubated eggs originating from Suriname (Godfrey 1997) and from Mona Island, Puerto Rico (Mrosovsky et al. 2009) to estimate leatherback hatchling sex ratio in JNP and hawksbill hatchling sex ratio in Saona respectively. To apply these curves to artificially incubated clutches, we added 4.1 days to our incubation duration data, which is the only published estima-tion of the time gap between hatching and emergence of hatchlings at the sand surface (Godfrey and Mrosovsky 1997).

3. 6. TRACkING HAWkSBILL TURTLES

Platform Terminal Transmitters (PTT) were attached on nesting hawksbill in Sao-na Island (n = 8) and JNP (n = 1). To pre-pare turtles for satellite tag attachment, a portable wooden corral was erected around each turtle following nesting. The carapace

Figure 3.10. Releasing leatherback and hawksbill turtle hatchlings in Bahía de las Águilas and Saona Island. (Photos courtesy of: R. Briones, Y.M. León, P. Feliz, J. Tomás and O. Revuelta).

3. 5. ESTIMATING HATCHLING SEx RATIO

Since the sex of marine turtle hatchlings can not be assessed from external mor-phology and scarifying hatchlings was not an option from an ethical and conservation perspective, incubation duration of clutches was used as an indirect method to estimate hatchling sex ratio (Mrosovsky et al. 1999). For each clutch, we calculated the incuba-tion duration (defined as the number of days between egg laying and the first hatch-ling emergence). This method has been devised for situations in which only pivotal incubation duration (that duration giving 50% of each sex) is available from labora-tory studies, and no field samples have been collected and sexed. The laboratory studies


within the polygon. Hence, core-use areas of activity were identified using fixed kernel density estimation (KDE) with individual kernel contours delineated using a smooth-ing factor (h) with the spatial analyst ex-tension of ArcGIS (ESRI). We used a 90% KDE to represent the overall home range of a turtle and a 50% KDE to represent the core area of activity (Powell 2000). To analyze the location of turtles with respect to marine protected areas (MPA), MPAs boundaries were overlaid on all resulting maps and summed location data with re-spect to the boundaries. Site fidelity for the females was quantified using a residency in-dex (Mason and Lowe 2010) calculated by dividing the number of days a female was detected within the DENP’s boundary by the number of days the female was moni-tored in the area (i.e. internesting period).

Bathymetric data were sampled from the General Bathymetric Chart of the Oceans GEBCO 1-Minute Global Bathymetry Grid (www.bodc.ac.uk/projects/interna-tional/gebco/gebco_digital_atlas). MPAs boundaries were downloaded from the World Database on Protected Areas (www.wdpa.org).

of each turtle was prepared by scrubbing to remove epibionts, sanding lightly, and cleaning with acetone, and a PTT was at-tached with 2-part epoxy (based on the methods of Blumenthal et al. 2006; Figure 3.11).

Turtle positions were determined with the ARGOS system which assigns location ac-curacy estimates (location class [LC]) to each reported location that are classified as 3, 2, 1, 0, A, B, or Z. Argos assigns accu-racy estimates of <250 m for LC 3, 250 to <500 m for LC2, 500 to <1500 m for LC1, and >1500 m for LC0 (CLS, 2011). The estimated accuracy is unknown for LCs A and B, and locations failing the Argos plausibility tests are tagged as class LCZ. Tracking and remote sensing data were downloaded and filtered using the Satellite Tracking and Analysis Tool (STAT) (Coyne and Godley 2005) program available from http://seaturtle.org. Locations that were used for assessing movements and delineat-ing core-use areas in this study were taken from Argos Location Classes 3, 2, 1, A and B (shown to be the best locations to describe hawksbill movements; Gaos et al. 2012). Bi-ologically unreasonable results of location points were also filtered, such as unlikely swimming speeds (>5 km h-1; see Luschi et al. 1998), turning behaviour (<25° turn an-gle) or erroneous locations on land.

Home range was estimated using the mini-mum convex polygons (MCP), a non-statis-tical measure which encapsulates the area used by an individual within a polygon formed by joining the outer-most sighting positions (Burt 1943). MCP is a simple cal-culation that allows for comparisons be-tween studies (Hooge et al. 1999), however is unable to define fine-scale movements


Figure 3.11. Attaching satellite transmitters on nesting hawksbill turtles at Saona Island. (Photos courtesy of: O. Revuelta, J. Blumenthal and H. González).

3. 7. STATISTICAL ANALYSESIn chapters 5 we analyzed data using Gen-eralized Linear Mixed Models (GLMMs) that were performed with the lme4 package (Bates et al. 2008) in R (version 2.14.0) and the graphical output was produced with the sciplot package (Morales 2012). In chapter 6, statistical analyses were carried out using the statistical package SPSS v17 (IBM). In chapter 7, analyses were performed using the lnme package (Pinheiro et al. 2011) in R (version 2.14.0) for the linear mixed mod-els. In both chapters, model selection was


based on ΔAIC values lower than 2, cal-culated as the difference between the AIC values for each model and the model with lowest AIC, and model weights (wi) (Burn-ham and Anderson 2002). Data of chap-ter 8 were mapped in ArcGIS 10.0 (ESRI, 2010). Specific analyses developed for each study, or other statistical programs used, will be indicated in each chapter. Statistical significance was set at P < 0.05, unless oth-erwise stated.

4. CHAPTER IV: PROTECTED AREAS HOST IMPORTANT REMNANTS OF MARINE TURTLE NESTING STOCkS IN THE DOMINICAN REPUBLIC

© P

ablo

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ChApTEr IV: PROTECTED AREAS HOST IMPORTANT REMNANTS OF MARINE TURTLE NESTING STOCkS IN THE DOMINICAN REPUBLIC

CHAPTER IV:

PROTECTED AREAS HOST IMPORTANT REMNANTS OF MARINE TURTLE NESTING STOCkS IN THE DOMINICAN REPUBLIC

Ohiana Revuelta1, Yolanda M. León2, 3, Pablo Feliz3, Brendan J. Godley4, Juan A. Raga1 and Jesús Tomás1, 4

__

Published in Oryx (2012) volume 46: 348-358.

__1 Cavanilles Institute of Biodiversity and Evolutionary Biology, University of Valencia, P.O. Box 22085,

E-46071 Valencia, Spain.

2 Instituto Tecnológico de Santo Domingo, Urb. Galá, Santo Domingo, Dominican Republic.

3 Grupo Jaragua, El Vergel, Santo Domingo, Dominican Republic.

4 Centre for Ecology and Conservation, School of Biosciences, University of Exeter Cornwall Campus, Penryn, UK.

04

Chapter IV: Protected areas host important remnants of marine turtle nesting stocks in the Dominican Republic l 91

ABSTRACTNesting by marine turtles in the Caribbean has declined considerably, mainly because of human exploitation, but there has previously been no monitoring in the Dominican Republic. We present the first detailed assessment of the status of marine turtle nesting in the country, based on surveys during 2006-2010. Nesting populations of hawksbill Eretmo-chelys imbricata and leatherback turtle Dermochelys coriacea are of regional importance and the green turtle Chelonia mydas is still present, although nesting in low numbers. The two main nesting sites are within protected areas: the Jaragua National Park in the south-west, important for leatherback turtles (mean of 126 nests per season), and Del Este National Park on Saona Island in the south-east, principally for hawksbill turtles (mean of 100 nests per season). Comparison with historical data suggests all rookeries are profoundly reduced in size. Although the main nesting beaches are within protected areas, illegal egg-take and meat consumption continues there and also elsewhere in the country.


and juvenile and adult green Chelonia mydas and hawksbill turtles Eretmochelys imbricata were exploited for their meat, eggs, shell and other products (Ottenwalder 1981; Dominguez and Villalba 1994; Mota and León 2003). By the 1980s it was estimated that 1,000-2,000 turtles were taken annual-ly, of which 70% were green and hawksbill turtles and 30% loggerhead Caretta caretta and leatherback turtles (Ottenwalder 1996). Most of these turtles were caught by divers or taken incidentally in seine nets but many were captured from beaches whilst nesting. Despite legislation banning trade in tor-toiseshell, an estimated 600 kg of hawksbill shell were used annually in the Dominican Republic in the 1980s (Ottenwalder 1987). Between 1970 and 1986, Japanese cus-toms data indicate that a total of 4,366 kg of hawksbill shell were exported to Japan (Fleming 2001). Domestically, sale of items made from tortoiseshell has been reported more recently (Mota and León 2003; Reu-ter and Allan 2006; Feliz et al. 2010).

The Dominican Republic receives four mil-lion visitors annually, and this mass tour-ism has resulted in degradation of coastal habitats (León 2004; Wielgus et al. 2010). Between 1980 and 1996 seven major coast-al areas, including a large proportion of the most important turtle nesting habitat, were developed (Ottenwalder 1996). By the 1980s six marine protected areas had been established, four of which (Montecristi, Del Este, La Caleta and Jaragua National Parks) cover c. 22% of the coastline. There are nine coastal parks managed by the Nation-al Parks Directorate and, although none of them were created specifically for protec-tion of marine turtles, some provide nesting and foraging habitat for them. However,

INTRODUCTIONMany marine turtle nesting rookeries in the Caribbean have been reduced or ex-tirpated by human exploitation (Parsons 1962; Bjorndal and Jackson 2003; Bell et al. 2006; Bräutigam and Eckert 2006). Al-though a number of rookeries have been studied for several decades and their sta-tus is well documented (Dutton et al. 2005; Troëng and Rankin 2005; Richardson et al. 2006; Diez and van Dam 2007; Beggs et al. 2007), many other Caribbean rookeries remain poorly described (McClenachan et al. 2006; Dow et al. 2007). The Dominican Republic is an area where information on marine turtle nesting activity is scarce and outdated. Unlike foraging habitats, which have been studied since 1996 (León and Diez 1999; León and Bjorndal 2002), the conservation status of nesting populations in the Dominican Republic has never been comprehensively assessed, although there have been suggestions that they are seri-ously threatened (Ottenwalder 1981, 1987; Stam and Stam 1992; Dow et al. 2007).

The Dominican Republic has a long histo-ry of harvest of and commerce in marine turtles, which have constituted an import-ant resource for coastal communities (Ot-tenwalder 1981; Fleming 2001; Mota and León 2003; Reuter and Allan 2006). Marine turtles are now legally protected and their trade banned in the country by laws dating from 1966 and recently confirmed through the 2005 Fisheries Law CODOPESCA (Dominican Council of Fisheries and Ag-riculture). Slaughter of leatherback turtles Dermochelys coriacea for their meat and eggs was widely recorded in the 1980s (Otten-walder 1981; Ross and Ottenwalder 1983)

Ohiana

Resaltado


MATERIALS AND METHODS

Study areaThe Dominican Republic is in the eastern part of the Caribbean island of Hispanio-la, which it shares with Haiti. It has 1,389 km of shoreline of which c. 800 km are sandy beaches. In February 2006 beach surveys and interviews with local people were carried out on 31 beaches previously described as important nesting sites (Ot-tenwalder 1981; Ross and Ottenwalder 1983). Following these surveys intensive sampling concentrated in the two areas where nesting still seemed to be significant: Jaragua National Park and Saona Island (Figure 4.1). Opportunistic surveys were conducted in other areas during 2006-2010 to detect any other potential nesting sites and to verify reports received during the study (Figure 4.2).

Jaragua National Park in the south-west covers 1,374 km2, of which 905 km2 com-prises beaches and dunes. The westernmost beaches, Bahía de las Águilas (4.4 km long) and La Cueva (2.5 km), have fine-grained, coralline, white sand. Although these beach-es have no infrastructure development, they receive >24,000 visitors per year (Wielgus et al. 2010). The easternmost beaches of San Luis (11 km), Mosquea (3.3 km) and Inglesa (1.2 km) are on the narrow area of land between Oviedo Lagoon and the Ca-ribbean Sea (Figure 4.1). Eastern beaches are wider than those in the west and have coarser, darker sand, a steeper slope and strong wave action, which contributes to the accumulation of large amounts of plas-tic and other debris on the beaches of San Luis and Mosquea (OR, YML, PF and JT,

there is little information available on ma-rine turtles nesting in these areas.

Surveys in the 1970s and 1980s previous-ly constituted the main reference on the status of marine turtles in the Dominican Republic (Ottenwalder 1981; Ross and Ot-tenwalder 1983). Based on a combination of five countrywide aerial surveys conduct-ed in April-July 1980, interviews with fish-erman and local people, and nonintensive beach surveys, Ottenwalder (1981) esti-mated 420 (range 240-600) hawksbill, 380 (range 253-507) leatherback, 260 (range 160-360) green and 60 (range 30-90) log-gerhead turtles nesting annually. Ross and Ottenwalder (1983) identified four areas of special interest for nesting leatherback turtles: the beaches of San Luis and Bahía de las Águilas in Jaragua National Park in the south-west and Del Muerto and Macao beaches on the east coast. More recently leatherback turtle nesting has been record-ed on the eastern beaches of Jaragua Na-tional Park during irregular beach walks during April-June (Dominici 1996).

The lack of comprehensive studies and re-cent information and the threatened status of marine turtles in the Dominican Repub-lic necessitated an updated assessment to help target effective conservation action. Our study had three main objectives: (1) to present the first systematic assessment in 30 years of the status of marine turtles nesting in the entire Dominican Republic based on 5 years (2006-2010) of systematic surveys, (2) to identify the main nesting rookeries and describe the current spatio-temporal patterns of nesting, and (3) to assess the likely impact of the current threats to these nesting stocks.


Mano Juan village and the four surveyed areas of (1) El Toro, (2) Mano Juan, (3) Canto la playa and (4) Faro Punta Cana. The insets indicate the location of the main maps in the Caribbean.

Data collectionDaytime surveys were undertaken on foot by researchers and trained local people on the beaches of Jaragua National Park and Saona during 2006-2010. In Jaragua Na-tional Park beaches were monitored during: (1) 15 April to 19 November 2006, (2) 17 March to 20 October 2007, (3) 7 March to 8 October 2008, (4) 14 March to 23 Sep-tember 2009, and (5) 26 March to 9 Sep-tember 2010. Until 2009 the eastern beach-es of the Park were monitored weekly, with the exception of the remote Inglesa beach, which was monitored twice per month. In 2009 and 2010 Mosquea was monitored daily by government rangers. On the west-ern beaches of the Park there were 3-4 day-time surveys per week but survey effort en-sured the record of all nesting events every year. Other remote beaches at this site with low levels of nesting were also occasionally visited to confirm reports of nests from re-liable informants, although these were few (16 clutches in 5 years). On these beaches successful nesting events and false crawls that had occurred since the previous visit were recorded.

On Saona Island surveys started in Septem-ber 2006 and thus we do not include total nesting numbers for this location and year. Since 23 September 2006 four major areas (Figure 4.1) were patrolled weekly through-out the year. Surveys in other areas of the Dominican Republic were undertaken during 2006-2010. Visits, sporadic patrols and interviews were made in 11 areas of the

pers. obs.). Six human settlements with a total population of c. 15,000 are located around the boundary of the Park.

Saona Island (a part of Del Este Nation-al Park) in the south-east has an area of 110 km2. The main nesting zones for ma-rine turtles include 12 narrow, white sand beaches (15 km long in total) interspersed with rocky areas on the south and west of the island (Figure 4.1). Access to some of these beaches by foot is difficult. There is one permanent human settlement on Sao-na, Mano Juan, with a population of c. 300. Both Saona Island and Jaragua National Park are regularly visited by groups of itin-erant fishermen for several days at a time.

Figure 4.1. The two main study areas of (a) Jaragua National Park, indicating the beaches of La Cueva and Bahía de las Águilas in the west and of San Luis, Mosquea and Inglesa in the east, and (b) Saona Island, indicating


tagged with large metal Inconel tags (Na-tional Brand and Tag Co., Newport, USA) fitted between the tail and rear flippers. Hawksbill females were tagged with small metal Inconel tags on the trailing edge of both fore-flippers.

RESULTS

Status and spatio-temporal distributionThe surveys confirmed that leatherback, hawksbill and green turtles are nesting in the Dominican Republic. Nesting is concentrat-ed in Jaragua National Park and on Saona Island. Nesting outside these areas was con-firmed, through visits and interviews, to be relatively uncommon (Figure 4.2).

The Jaragua National Park consistently had the highest number of clutches per year of leatherback turtles (mean 126.4 ± SD 74.1, range 17-210), with a total of 632 clutches recorded during 2006-2010 (Figures 4.2 and 4.3). These values correspond to an estimat-ed annual number of three, 33, 19, 40 and 25 females in each of the 5 years, respective-ly (based on a mean number of clutches per season of 5.26; Boulon et al. 1996). Hawks-bill turtles nested in low numbers in the Park (mean 14.6 ± SD 6.7 per year, range 7-22 clutches per year), with a total of 73 clutches recorded in the 5 years, mainly in the west (Figure 4.2). The estimated annual number of hawksbill turtles was five, four, four, one and two nesting females in each of the 5 years, respectively (based on a mean number of clutches per season of 4.5; Richardson et al. 1999). We recorded only one confirmed clutch of green turtles in this area during the study period (Figure 4.2).

northern coast in 2006, 2007 and 2008, and on six beaches of the eastern and southern coasts in 2008, 2009 and 2010 (Figure 4.2). Reports about other nesting events, partic-ularly from resorts or beach front develop-ments, were also compiled.

During daytime surveys in Jaragua Nation-al Park and on Saona Island we recorded total number of emergences of nesting fe-males. We identified the species based upon track characteristics (Schroeder and Mur-phy 1999) and confirmed whether each activity had successfully resulted in clutch deposition, by digging to find the clutch. Nest location was recorded using a global positioning system and turtle tracks were marked to avoid duplicate counts. Signs of eggs taken by people were recorded (pres-ence of probing sticks, evidence of digging and broken eggshells, and/or human foot-prints around the nesting site; Troëng et al. 2004). Remaining clutches were either cam-ouflaged and incubated under natural con-ditions on the beach, or clutches with high risk of subsequent take (because of prox-imity to a human settlement and the num-ber of people in the area) were transferred to protected hatcheries for incubation in boxes. Because of funding limitations night surveys were only carried out on Saona Is-land and the western beaches of Jaragua National Park. To maximize the probabil-ity of finding nesting females during night surveys, we timed these surveys according to the inter-nesting intervals of the females. We collected data on clutch size and egg size and weight and measured the curved carapace length (CCL, from nuchal notch to the caudal tip of the carapace) and maxi-mum curved carapace width (CCWmax) of nesting females. Leatherback females were


Elsewhere 1-5 hawksbill turtle nests per year were reported on the Punta Cana and Cap Cana beach resort on the east coast. The only beach outside the two main nesting areas that seems to host significant leath-erback turtle nesting is El Muerto beach (Figure 4.2). In 2009 we recorded 12 recent nests there. Local people reported nine nests later in the same season. This area merits future monitoring. The northern coast of Samaná Peninsula seems regularly to have a low number of nests (1-3 clutches per beach and year) of hawksbill and leath-erback turtles, and merits further surveys.

Saona Island hosted the majority of re-corded hawksbill (mean 100 ± SD 8.4 per year, range 93-111, total 400) and green turtle clutches (mean 9.2 ± SD 6.2, range 1-15, total 37) but only 22 leatherback tur-tle clutches were recorded (mean 5.5 ± SD 4.8, range 1-11). The annual number of hawksbill turtle clutches recorded was sim-ilar among years (Figure 4.3) and the esti-mated annual number of hawksbill females nesting in each year was 21, 23, 23 and 25 in 2007-2010, respectively.

Figure 4.2. Maximum number of confirmed nests of hawksbill Eretmochelys imbricata, leatherback Dermochelys coriacea and green Chelonia mydas marine turtles per year during 2006-2010. Numbers correspond to beach locations: (1) beaches on both sides of the Estero Hondo bay (Punta Rucia and Punta Burén), (2) Sosúa and Cabarete beaches (Puerto Plata province), (3) Playa Grande, (4) Arroyo Salado, (5) Cosón (bonita and Morón beaches), (6) El Limón, (7) Lanza del Norte, (8) El Valle, (9) Rincón, (10) Colorá, (11) Las Galeras/Madama, (12) El Muerto/La Vacama, (13) Macao and Uvero Alto, (14) Punta Cana, (15) southern Saona Island (several beaches), (16) Isla Catalina, (17) San Pedro de Macorís, (18) Guibia (Santo Domingo), (19) Salinas, (20) Oviedo Lagoon, (21) Beata Channel, and (22) La Cueva-Bahía. Dotted lines show the two main study areas: Jaragua National Park in the south-west and Saona Island in the south-east (part of Del Este National Park).


clear because of the low number of nests observed. The three species nested mainly at night although there were two records of hawksbill turtles nesting diurnally on Saona Island.

Figure 4.4. Mean (± SD) number of clutches per month of leatherback (black) and hawksbill (white) marine turtles for 2006-2010 in Jaragua National Park, and for 2007-2010 on Saona Island. Note the different y-axis scales.

Morphometrics and reproductive dataBiometric parameters of leatherback and hawksbill turtles are given in Table 4.1. All leatherback females encountered on western beaches of Jaragua National Park in 2006-2010 were tagged. We recorded four reproductive females (30.8%) with a CCL (132, 135, 138 and 143 cm) below the threshold carapace length for adult classifi-cation in leatherback turtles (145 cm; Eck-ert 2002; James et al. 2007). Mean clutch sizes (yolked eggs) of leatherback turtles re-

Figure 4.3. Number of clutches of leatherback (black), hawksbill (white) and green (grey) marine turtles per year in Jaragua National Park and on Saona Island. In 2007 one green turtle nest, which is not shown, was recorded in Jaragua National Park. Note the different y-axis scales.

SeasonalityIn Jaragua National Park the nesting pe-riod of leatherback turtles extended from March to August, with most emergences occurring in April-June (89.6% of nests). The highest number of emergences was in May (mean number of nests 64 ± SD 12.7). Because of the low level of nesting by hawksbill turtles in the Park it is difficult to ascertain seasonality for this species there, although nesting appears to be higher in July (Figure 4.4). On Saona Island hawks-bill turtles were observed nesting in every month of the year, although most nesting occurred in June-November (71.2% of the nests) and low levels of nesting were ob-served in December-May (Figure 4.4). The green turtle appears to have a shorter nest-ing season, from July to November, with a possible peak in August. However, any sea-sonality in the nesting of this species is un-


and only a small percentage of leatherback clutches were incubated on the beach (1% in 2007 and 8.4% in 2009); the remainder were relocated to a protected hatchery for artificial incubation or otherwise taken by humans. The percentage of leatherback turtle clutches taken on the beaches of Bahía de las Águilas and La Cueva in the west of the Park was lower than on the east-ern beaches, with 75.9% of 290 clutches in-cubated naturally during the study period. However, take of hawksbill turtle clutches was high every year (Table 4.2).

We found hawksbill and green turtle car-casses, bones and scutes from both adults and juveniles in Jaragua National Park and on Saona Island, which we presume to be remains, at least in part, from illegal take. Meat and eggs are consumed and illegal-ly sold in local markets (YML and JT, pers. obs.). Throughout the study period no re-ported cases of consumption of leather-back meat were recorded.

corded in the Park were 68.7 ± SD 18.1 (n = 64), 72.4 ± SD 15.5 (n = 59), 75.1 ± SD 17.3 (n = 125) and 67.5 ± SD 17.8 eggs (n = 67) in 2007-2010, respectively.

On Saona Island seven female hawksbill turtles were tagged in 2008, five in 2009 and seven in 2010 (Table 4.1). Mean clutch sizes of hawksbill turtles on Saona Island were 125.7 ± SD 23.3 (n = 29), 132.4 ± SD 29.9 (n = 55), 138.4 ± SD 23.2 (n = 40) and 139.5 ± SD 29 (n = 55) in 2007-2010, respectively.

Assessment of threatsIllegal taking of eggs was identified as the main threat to the marine turtles (Table 4.2). The highest level of egg-take was from the eastern beaches of Jaragua Nation-al Park. All clutches laid by hawksbill and leatherback turtles on Inglesa beach were taken by humans. Overall 88.9% of hawks-bill clutches laid on Mosquea and San Luis beaches during the study period were taken

Table 4.1. Morphometric and reproductive parameters of leatherback (Dermochelys coriacea) and hawksbill (Eretmochelys imbricata) turtles nesting at Jaragua National Park and Saona Island in the Dominican Republic (Figure 4.1).

Leatherback (jNP) Hawksbill (Saona)

N Mean ± SD Range N Mean

± SD Range

Curved carapace length (cm) 13 147.4 ± 8.7 132-162 19 87.2 ± 4.7 81-96

Curved carapace width (cm) 13 107.7 ± 5.1 100-118 15 77.1 ± 4.2 71-84

Clutch size (no. of yolked eggs) 315 71.7 ± 17.5 10-128 179 134.8 ± 27.5 82-235

Egg size (mm) 740 5.2 ± 0.2 4.5-5.7 380 3.4 ± 0.09 3.1-3.8

Egg mass (g) 500 85.4 ± 8.7 65.5-102.4 240 29.9 ± 2.3 24.5-85.4


Table 4.2. Total numbers and percentage of hawksbill, leatherback and green (Chelonia mydas) turtle clutches recorded in Jaragua National Park and Saona Island during 2006-2010 that were subject to human take, translocated to a breeding facility, or incubated on the beach under natural conditions.

Leatherback Hawksbill Green

Year (by

location)Total Human take

(%)Translocated

(%)Beach

(%) Total Human take(%)

Translocated(%)

Beach(%) Total Human take

(%)Translocated

(%)Beach

(%)

JARAGUA NATIONAL PARk (WEST)

2006 2 50.0 0.0 50.0 22 68.1 27.2 4.5

2007 59 37.3 10.2 52.5 14 71.4 28.6 0.0 1 0.0 100.0 0

2008 63 1.6 34.9 63.5 15 33.3 66.7 0.0

2009 83 4.8 16.9 78.3 4 25.0 50.0 25.0

2010 83 0.0 0.0 100.0 8 37.5 12.5 50.0

JARAGUA NATIONAL PARk (EAST)

Mosquea-San Luis

2006 7 42.8 57.1 0.0

2007 102 48.0 51.0 1.0 4 75.0 0.0 25.0

2008 29 44.8 55.2 0.0 2 100.0 0.0 0.0

2009 119 23.5 68.1 8.4 3 100.0 0.0 0.0

2010 39 46.1 51.3 2.6

Inglesa

2006 8 100.0 0.0 0.0 1 100.0 0.0 0.0

2007 12 100.0 0.0 0.0

2008 6 100.0 0.0 0.0

2009 8 100.0 0.0 0.0

2010 12 100.0 0.0 0.0

SAONA

2007 1 100.0 0.0 0.0 94 56.4 21.3 22.3 8 37.5 37.5 25.0

2008 2 50.0 50.0 0.0 102 32.3 50.0 17.6 15 13.3 73.3 13.3

2009 11 72.7 9.1 18.2 93 50.5 41.9 7.5 1 0.0 100.0 0.0

2010 8 75.0 25.0 0.0 111 17.2 46.8 36.0 13 7.7 38.5 53.8


Firstly, data reported by Ottenwalder (1981) were based on 2 years of non-inten-sive monitoring and interviews. Secondly, loggerhead turtles are considered to nest in-frequently in the Caribbean (Ehrhart et al. 2003; Dow et al. 2007), and Ottenwalder noted that, in interviews, many fishermen confused this species with others. It is there-fore possible that the number of logger-head turtles nesting was previously overes-timated, as is thought to have occurred in similar studies based on interview data in the region (Richardson et al. 2009). Thirdly, the decline of the leatherback turtle nesting rookeries in the Dominican Republic could be because of a shift of nesting to nearby rookeries, as this species has low nest site fidelity (Georges et al. 2007; Troëng et al. 2007). The leatherback turtle population in the Dominican Republic possibly forms part of a wider regional nesting stock, as has been suggested for other rookeries in the Antilles (Dutton et al. 2005). Despite these caveats it is probable that harvesting of eggs and females from beaches and the trade in marine turtles documented in the Dominican Republic (Ottenwalder 1996; Fleming 2001; Reuter and Allan 2006) has contributed significantly to the decline of these turtle populations. However, we can-not refute the possibility that numbers of all three turtle species nesting in the Domini-can Republic, particularly the leatherback and hawksbill, were lower in previous years and are beginning, as in some Caribbean range states, to increase (Dutton et al. 2005; Troëng and Rankin 2005; Richardson et al. 2006).

According to Ottenwalder (1981) hawksbill turtles nested in the past in significant num-bers on beaches around the entire coast of

DISCUSSIONFor the first time in recent decades we have documented the presence of three marine turtle species (hawksbill, leatherback and green) nesting in the Dominican Repub-lic. Comparison of our results with early reports (Ottenwalder 1981; Ross and Ot-tenwalder 1983) indicates that a profound decline appears to have taken place in the last 30 years. Nesting has largely been re-duced to the undeveloped, protected areas of Jaragua National Park and Saona Island. However, there are still high rates of human exploitation of eggs and at least some take of turtles in these protected areas.

The number of clutches of hawksbill and green turtles, which we found mainly on Sa-ona Island, are low compared to estimates of 400 hawksbill and 260 green turtles nesting per year throughout the country in the 1980s (Ottenwalder 1981). Ross and Ottenwalder (1983) suggested that 300 leatherback turtles were nesting annually in the Dominican Re-public, based on one aerial survey, 24 inter-views and a few beach surveys between 24 March and 13 April 1980. Based on 5 years of systematic surveys we have estimated that a maximum of 40 leatherback females cur-rently nest in Jaragua National Park per year, with no more than 50 in the country. Of fur-ther concern was our failure to detect log-gerhead turtles nesting, even though Otten-walder (1981) estimated c. 60 females were nesting per year.

Our findings seem to indicate that there has been a marked reduction in the abundance of the four species throughout the country since the 1980s. However, there are some caveats regarding the previous estimates.


1996; Mota and León 2003; Reuter and Al-lan 2006; Feliz et al. 2010). However, since 2009 the environmental authorities have conducted several seizures of tortoiseshell items in the gift shops of Santo Domingo, resulting in a significant reduction in the sale of such items (OR, YML and PF, pers. obs.).

Our study has shown that the leatherback turtle nesting season is from March to Au-gust, with a peak in May, consistent with re-ports for neighbouring nesting populations (Dutton et al. 2005; Hilterman and Gov-erse 2007; McGowan et al. 2008). The sea-sonality of hawksbill turtle nesting is similar to nearby rookeries on Antigua, Barbados and Mona Island, Puerto Rico (Richardson et al. 1999; Beggs et al. 2007; Diez and van Dam 2007).

The differences in nesting activity between the two main nesting areas may be explained by species’ preferences. Leatherback turtles tend to use high energy beaches with deep water nearby, free of obstructions and often with a windward orientation (Mrosovsky 1983; Eckert 1987). This description fits the eastern beaches of Jaragua National Park, which is the most important nesting area for the species in the country. Hawksbill turtles, however, often swim over reefs and through shallow waters with lower wave energy to reach beaches with dense vegetation (Hor-rocks and Scott 1991). This description fits all of the Saona Island beaches and most of the western beaches of Jaragua National Park.

The mean CCL (147.4 cm) of leatherback turtles in the Dominican Republic is low-er than the global mean, which is normally distributed around 155 cm (see review in

the Dominican Republic and green turtles nested every night on northern beach-es. Currently, only the relatively unspoilt beaches of Saona Island, in Del Este Na-tional Park, provide major nesting areas for hawksbill and green turtles. Nesting of leatherback turtles previously occurred on the northern and eastern coasts, and the beaches of Macao and El Muerto in La Al-tagracia province on the eastern coast host-ed 2-3 females nightly during the nesting season in the 1980s (Ross and Ottenwalder 1983). Significant nesting by leatherback turtles occurs only on the isolated beach-es of Jaragua National Park, where access for people is difficult and there is no coastal development. In addition to the taking of eggs and adults on beaches near habitation, habitat has been lost to intensive coastal development during the last 30 years (Ot-tenwalder 1981; León 2004; Wielgus et al. 2010). Currently, the northern and eastern coasts have the most tourism infrastruc-ture and we detected only sporadic nesting events on these beaches.

Despite the importance of coastal and ma-rine protected areas as refuges for marine biodiversity in the country we observed that there is little actual protection and in-adequate management by local authorities (OR, YML, PF and JT, pers. obs.). Our re-sults highlight the need for adequate pro-tection and management of these areas for marine turtle conservation, especially considering that Jaragua National Park also hosts important foraging habitats for Carib-bean hawksbill and green turtles (León and Diez 1999; León and Bjorndal 2002).

Items made from tortoiseshell were previ-ously widely available in tourist gift shops in the Dominican Republic (Ottenwalder


this is the case in the Dominican Republic, which therefore warrants increased atten-tion at the regional level, particularly for the hawksbill turtle. Molecular and genetic techniques have provided insights into pat-terns of migration, stock differentiations, and links between nesting beaches and for-aging grounds of the hawksbill turtle in the Caribbean (Bowen et al. 2007; Blumenthal et al. 2009). Recovery will therefore depend on actions both in the Dominican Republic and in the wider Caribbean.

The fact that nesting activity is concentrated in protected areas provides an opportunity to implement specific actions for protection and management of the rookeries (San-tidrián Tomillo et al. 2008). We conclude that to improve marine turtle conservation in the Dominican Republic it is necessary to promote the effective management of existing protected areas, including nesting beaches for marine turtles, to promote the management of coastal development to avoid further loss of nesting beaches, and to support the enforcement and efficient implementation of laws to reduce take of marine turtles and their eggs.

ACkNOWLEDGEMENTSThe present study is part of a conservation project funded by the Spanish Internation-al Cooperation Agency (AECI, projects A/2991/05 and A/5641/06), the Span-ish Ministry of Education and Sciences (CGL2006-02936-BOS and CGL2011-30413), the General Foundation of the Uni-versity of Valencia, a special action of the same University (UVINV- AE11-42960), and the European Union (Marie Curie

Stewart et al. 2007) and we confirmed the viability of clutches of females that are be-low the minimum size of sexual maturity reported elsewhere (e.g. Eckert 2002). We are confident that our results are not mea-surement errors because we used standard measurements (Bolten 1999) and the cara-paces had no deformities or injuries. This could be the sign of a recovering popula-tion with a high proportion of neophyte females. However, reduction in the size of nesting females can be caused by intensive capture of turtles either in the foraging ar-eas or in the same nesting areas (Bjorndal et al. 1985; Limpus et al. 2003). The mean CCL of nesting hawksbill turtles in the Do-minican Republic is similar to the mean CCL in other Caribbean regions. However, the CCL of 11 hawksbill females measured on Saona Island was within the CCL range recorded for neophytes in Buck Island, US Virgin Islands (82.1-88.0 cm, Garland and Hillis-Starr 2003).

This research has filled a significant gap in knowledge of the population status and conservation of marine turtles in the Carib-bean (Dow et al. 2007). Many of the large turtle rookeries in the region are the sub-ject of ongoing long-term monitoring and conservation programmes. Although sever-al appear to be showing signs of recovery (St Croix, US Virgin Islands, Dutton et al. 2005; Barbados, Beggs et al. 2007; Mona Island in Puerto Rico, Diez and van Dam 2007) this is not the case for all turtle rook-eries in the region.

To facilitate effective marine turtle conser-vation across the Caribbean conservation programmes need to focus on sites that were formerly important nesting areas (Mc-Clenachan et al. 2006). Our results suggest


grants, FP6 and 7). We thank the Domini-can NGO Grupo Jaragua and The Nature Conservancy. The project counts on the support of the Dominican Environmental authorities (Office of Protected Areas and Biodiversity of the Environmental Minis-try) and the rangers of the National Parks. We are grateful to our assistants in Jaragua National Park, Bienvenido Pérez Turbí ‘Blanco’ and the rangers for their work, and especially to Pelagio Paulino ‘Negro’ for his help during fieldwork on Saona Island and his work for the conservation of marine tur-tles. We extend our thanks to the Domini-can and international students, volunteers and villagers for their help and friendship during the field work.


Chelonia, Inc., and Universidad de Puerto Rico, Rio Piedras, Puerto Rico.

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5. CHAPTER V: ASSESSING THE EFFICACY OF DIRECT CONSERVATION INTERVENTIONS: CLUTCH PROTECTION OF THE LEATHERBACk MARINE TURTLE IN THE DOMINICAN REPUBLIC

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ChApTEr V: ASSESSING THE EFFICACY OF DIRECT CONSERVATION INTERVENTIONS: CLUTCH PROTECTION OF THE LEATHERBACk MARINE TURTLE IN THE DOMINICAN REPUBLIC

CHAPTER V:

ASSESSING THE EFFICACY OF DIRECT CONSERVATION INTERVENTIONS: CLUTCH PROTECTION OF THE LEATHERBACk MARINE TURTLE IN THE DOMINICAN REPUBLIC

Ohiana Revuelta1, Yolanda M. León2, 3, Annette C. Broderick4, Pablo Feliz3, Brendan J. Godley4, Juan A. Balbuena1, Andrea Mason4, Kate Poulton4, Stefania Savoré4, Juan A. Raga1 and Jesús Tomás1, 4

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Oryx (In Press)





4 Centre for Ecology and Conservation, School of Biosciences, University of Exeter Cornwall Campus, Penryn, UK.

05

Chapter V: Assessing the efficacy of direct conservation interventions: clutch protection of the leatherback marine turtle in the Dominican Republic I 113

ABSTRACTIn the Dominican Republic (DR), northern Caribbean, the beaches of the Jaragua Na-tional Park (JNP, southwest of DR) are the last major nesting site for the leatherback ma-rine turtle, Dermochelys coriacea, in the country (126.4 ± SD 74.1, clutches per year, range (17-210): 2006-2010). This nesting aggregation is highly threatened due to a widespread illegal egg take. Clutch relocation and artificial incubation have been carried out here as a protection measure since 1974. We sought to critically assess the efficacy of such efforts and investigated how artificial incubation may be influencing success of clutches and re-sultant sex ratios. We compared hatching success, incubation duration and embryo mor-tality from in situ clutches (n = 43 clutches) with those artificially incubated at sites in the east and west of the park (n = 35 and n = 31 clutches, respectively). Our study found that in the west artificial incubation significantly decreased hatching success in clutches and, in the east, increased incubation duration, which we predict would result in an increase in male production from these clutches. Clutch relocation is currently the only viable conser-vation option for clutches on eastern beaches due to intense egg take, but steps are needed to ensure that natural sex ratio is not distorted. However, on the western beaches, in situ clutch incubation seems possible through beach protection. Further community engage-ment and enforcement is required to improve conservation measures at eastern beaches if long-term, less sustainable, intervention is to be avoided.


INTRODUCTIONIn critical conservation situations includ-ing endangered species urgent actions are demanded to preserve species and ecosys-tems (Andrews 2000). However, in many situations conservation actions have been passed on from individual to individual with no assessment to see whether or not these management actions fulfill the conservation objectives (Pullin and Knight 2003). Carry-ing out evaluations of these activities is nec-essary to demonstrate that conservation ac-tions actually achieve their objectives (Pullin and Knight 2009; Sutherland et al. 2009).

As a result of centuries of exploitation ma-rine turtles are recognised internationally as species of conservation concern (Hamman et al. 2010). The take of eggs by coastal people, once widespread around the world, is still a significant threat to the survival of some marine turtle populations (Santid-rián Tomillo et al. 2008). This has led to a variety of conservation measures, includ-ing protection of nesting beaches (Frazier 2002; Santidrián Tomillo et al. 2009), and programs of egg relocation from threat-ened sites to other beach locations or to en-closed hatcheries (Tuttle and Rostal 2010; Liles et al. 2011).

On nesting beaches where clutches are handled, management techniques carried out may affect offspring production (Pintus et al. 2009). The handling and transporta-tion of eggs may result in embryo death as rotation or vibration can rupture the em-bryonic membrane (Phillott and Parmenter 2007), resulting in a reduced hatching suc-cess of relocated clutches when compared to those incubated in situ (Eckert and Eck-

ert 1990; Özdemir and Türkozan 2006). Temperature, humidity and oxygen levels in relocated clutches are likely to be different from those chosen by the nesting female, therefore hatching success, sex ratios and phenotype may be affected (Foley et al. 2000). In marine turtles, the temperature experienced by the embryo during the middle third of development determines the sex of the hatchling (Yntema and Mrosovsky 1982) and while female biased offspring sex ratios appear to be the norm for marine turtles, there have been exam-ples of artificial incubation carried out un-der cooler thermal conditions resulting in male biased offspring sex ratios (Mrosovsky 1982; Morreale et al. 1982; Dutton et al. 1985). Furthermore, temperature may also affect the embryonic development with very high or very low incubation temperatures increasing embryo mortality rates (Acker-man 1997; Broderick et al. 2001).

The leatherback marine turtle, Dermochelys coriacea, is currently listed as critically en-dangered by the world conservation union (IUCN 2012). However, the population sta-tus of this species varies among locations. Leatherback nesting populations in the In-do-Pacific have declined precipitously in recent decades (Spotila et al. 2000; Sarti Martínez et al. 2007) whilst many Atlantic nesting populations are stable or increasing (Dutton et al. 2005; McGowan et al. 2008; Witt et al. 2011). The Wider Caribbean re-gion holds some of the globally important nesting sites, such as Trinidad and Tobago (Dow and Eckert 2011) and the Caribbe-an coast of Colombia and Panama (Pa-tiño-Martinez et al. 2008). However, there are many small, widely dispersed sites that may play an important role in species re-


covery that lack intensive population moni-toring effort (Dow et al. 2007).

In the Dominican Republic, the leather-back turtle nests sporadically around the coast, but the beaches of the Jaragua Na-tional Park (JNP, southwest of DR) have been recently identified as the last import-ant stable nesting site for the species in the country, (mean of 126.4 clutch.year-1 SD 74.1, range 17-210 over 5 seasons; 2006 to 2010; Revuelta et al. 2012). Recent analyses with genetic markers show that leatherback turtles in the Dominican Republic appear to be connected with many other Carib-bean populations (Carreras et al. 2013). Although the harvesting of turtles or their products has been banned since 1966, the enforcement of environmental law in the DR is weak and the leatherback nesting stock of JNP is highly threatened due to a widespread illegal egg take (Revuelta et al. 2012). However, there is marked spatial het-erogeneity in the level of human predation; illegal egg take is close to 100% in the east-ern beaches, whilst egg take is much low-er at western beaches (75.9% of clutches could be incubated in situ 2006 to 2010; Re-vuelta et al. 2012). In the face of this pres-sure there is a 38 year history of artificial incubation in JNP. In 1974, a local assistant was trained to carry out relocation and arti-ficial incubation of eggs in Styrofoam boxes (Ottenwalder 1981), placing them in a facil-ity located 10 km from the nesting beach. However, this activity has been carried out sporadically with no strict protocol or sci-entific monitoring. In 2006 we initiated a project to assess this conservation program investigating how artificial incubation may be influencing hatching success of clutches and resultant sex ratios.


Study SiteJaragua National Park is a protected area of 1374 km2 (of which 905 km2 is marine reserve) situated in the south western cor-ner of the DR (Figure 5.1). There are five main turtle nesting beaches grouped in two areas: (1) Bahía de las Águilas (4.4 km) and La Cueva beach (2.5 km) (western beach-es [WB], N17° 57’, W71° 39’), and (2) Mosquea beach (3.3 km), San Luis beach (11 km) and Inglesa beach (1.2 km) (eastern beaches [EB], N17°44’, W71° 20’) (Figure 5.1). The two areas are separated by ap-proximately 50 km. The western beaches border calm shallow waters and have fine-grained, white coralline sand (width ranges: 4-20 m) backed by low scrub vegetation. A small village (around 20 inhabitants) is lo-cated between the two beaches, which, in addition, are frequented by several thou-sand tourists per annum.

The eastern beaches (EB) lay beyond the Oviedo Lagoon, and are generally wider than the western ones (width ranges: 7-40 m), and have coarser, yellow sand, and steeper beach slope. They are exposed to the prevailing north-easterly winds and currents which deposit large amounts of plastic and other debris. Although not fre-quented by tourists, these eastern beaches are patrolled by beachcombers, some of whom partake in a high level of illegal egg take for consumption and sale (Revuelta et al. 2012).


Figure 5.1. Map of JNP, indicating the beaches of La Cueva and Bahía de las Águilas (western beaches) and of San Luis, Mosquea and Inglesa (eastern beaches).

SurveysIn 2006 and 2007, we carried out prelim-inary weekly beach surveys on the west-ern and eastern beaches of JNP to obtain baseline data on marine turtle nesting ac-tivity. Based on the observed differences in human egg take between western and eastern beaches and the availability of hu-man resources, we applied different survey and conservation strategies in each area of study in subsequent years. Although some artificial incubation was carried out in these two seasons, it was not until 2008 that we were able to test robustly for effects of arti-ficial incubation interventions.

Eastern beaches: In 2008 we carried out daytime weekly surveys during the leath-erback nesting season (from March until August, Revuelta et al. 2012) with the ex-ception of the remote Inglesa beach which was monitored twice a month. In 2009, added to our weekly surveys, governmental rangers performed nightly beach surveys in Mosquea and San Luis beaches and week-ly surveys in Inglesa beach. Owing to the high level of egg take at these beaches the

strategy chosen was the relocation and ar-tificial incubation of all clutches that were not-predated at the time of recording.

Western beaches: In 2008 and 2009, re-searchers and governmental rangers car-ried out daily surveys from March to Au-gust. Night patrols were also carried out two or three nights per week during the busiest part of the laying season (April and May, Revuelta et al. 2012) in order to witness laying events. These beaches are rarely visited by individuals involved in il-legal activities and access is controlled by rangers. For this reason, we left 50% (n = 22 clutches) and 76.9% (n = 40 clutches) of clutches incubating in situ in 2008 and 2009, respectively. In these years we limit-ed relocation to clutches that were thought unlikely to hatch without intervention (i.e., those clutches laid in a sand road near the beach or clutches threatened by tidal inun-dation). From 2010 onwards, we left 100% of clutches incubating in situ.

From 2008 we relocated clutches for artifi-cial incubation into the park rangers’ bar-racks. The facilities were a similar size and had the same characteristics at both areas: wooden barracks with concrete floor and roof made from corrugated metal panels. The WB barracks is on a hilltop about 2 km away from the beaches, and it is only acces-sible via a dirt track. This facility is located in an area of arid scrub and is exposed to the sun all day and to dry winds. Precipita-tion is more frequent at EB and the rangers’ barracks lies directly behind the Mosquea beach, inside a forest and near the Oviedo lagoon (See Figure 5.1).


Egg relocation and incubationDuring daytime surveys, clutches were lo-cated by careful probing of the sand with a stick at nesting sites discovered from tracks (Schroeder and Murphy 1999). During night surveys eggs were collected after the female had returned to the sea.

For artificial incubation we carefully exca-vated the egg chamber by hand and trans-ferred the yolked eggs into polyethylene exterior boxes with polyurethane foam filling (dimensions 30 cm width x 50 cm length x 32 cm depth) for their transport and subsequent incubation. Since artificial incubation at site before the project did not include the collection of yolkless eggs for in-cubation, and since a real function related to optimizing the clutch environment has not been proved (Wallace et al. 2006 and 2007) we decided to keep the same practice excluding the yolkless eggs from the boxes.

We tried to mimic the natural clutch ar-rangement as much as possible placing yolked eggs in the boxes using the sand from the original nesting beach. Two to three centimetres of sand were placed at the bottom and sides of the box to prevent contact of eggs with box walls, as well as over the eggs at the top of the box.

To record hourly clutch temperatures, we placed TinyTag Plus-2 dataloggers (Gemini Data Loggers UK Ltd., Model TGP-4017, accurate to ± 0.3°C) in artificially incubat-ed clutches in EB (n = 12) and WB (n = 33). An additional two loggers were placed as controls at nest depth in Mosquea and Bahía de las Águilas beaches in 2008 and 2009 but these were, unfortunately, lost to beach erosion.

When clutches were left in situ, we camou-flaged the tracks of females and recorded GPS coordinates. We recorded distance from nest to the high tide line, and the beach zone where the clutch was located (open beach, at the vegetation border or within the dune vegetation). Once the clutches were cam-ouflaged we estimated the date of hatching based on the average known incubation du-ration estimated of leatherback marine tur-tles (~ 60 days), and from 10 days before this date we checked the nest place in order to find hatchlings tracks.

Study of clutchesDuring the incubation, each box was checked daily for signs of hatched tur-tles. When artificially incubated clutches hatched, we carefully excavated them no earlier than 48 hours after the last sign of hatchling emergence. We excavated in situ clutches by hand one to three days after hatchlings emergence was detected in sur-veys. In both cases, data were collected on the number of hatched shells, live and dead hatchlings within the clutch. Clutch size was defined as the total number of eggs (hatched and unhatched). Hatching success was calculated by dividing the number of eggs hatched by the clutch size expressed as percentage (Miller 1999). We opened and examined the contents of all unhatched eggs. Eggs that had partial calcification of the shell or evidence of an early stage embryo were classified as early stage em-bryonic death (Bell et al. 2003). Eggs that contained dead hatchlings at a late stage of development were classified as late stage embryonic death.

For each clutch we calculated incubation duration (defined as the number of days


between egg laying and the time of first hatchling emergence). Incubation was considered completed when hatchlings emerged on the top of the sand. A regres-sion test was carried out to describe the re-lationship between incubation duration and mean temperature during the mid-third of incubation which is thought to contain the thermosensitive period (the time span outside of which temperature manipula-tions do not exert any influence on sexual phenotype, Mrosovsky and Pieau 1991). We estimated hatchling sex ratios using the conversion curve relating incubation dura-tion to hatchling sex ratio derived from in situ incubated eggs originating from Surina-me (Godfrey 1997). To apply this (Godfrey 1997) curve for artificially incubated clutch-es, we added 4.1 days to our incubation du-ration data, (Godfrey and Mrosovsky 1997). Examples of the use of this method and its validation are given in Zbinden et al. (2007) and Katselidis et al. (2012).

Statistical analysisTo understand the effects of incubation methods and nest metrics, we analyzed data using Generalized Linear Mixed Models (GLMMs) which allow both fixed and random factors as well as covariates to be fitted, and random factors control for the use of repeated measurements (Schall 1991). Our data were subject to temporal pseudoreplication caused by repeated mea-surements through time of the same fe-males nesting over the season. We therefore included clutch and year as random effects in the models to account for the pseudorep-lication among females and potential vari-ation in the response variable across years. GLMMs were performed with the lme4

package (Bates et al. 2008) in R (version 2.14.0) and the graphical output was pro-duced with the sciplot package (Morales 2012).

We examined the effect of incubation meth-od (artificial incubation at eastern beaches (EB), artificial incubation at western beach-es (WB) and in situ) on hatching success us-ing GLMM analyses with a binomial error, the most appropriate error distribution for percentages data, with a logit link function. Random effects considered were year and clutch and the covariates included were: incubation method, incubation duration (days), clutch size (number of yolked eggs) and laying date.

We investigated how incubation method affected the frequency of late and early stage dead embryos within studied clutch-es. We used subsequent GLMMs with Pois-son error, the most appropriate distribution for count data, and a log link function to examine the effect of incubation method on number of early stage dead embryos and late stage dead embryos occurring in clutches as in Pintus et al. (2009). Maximal models included clutch size (yolked eggs), incubation duration (days) and laying date as covariates as well as incubation method.

To analyze the effect of incubation meth-od on incubation duration we performed GLMM with a Poisson error and a log-link function; with the maximal model also in-cluding incubation method, laying date and clutch size as covariates.

For each analysis we started from a maximal model with all fixed effects and interactions. Significance of fixed effects was assessed us-ing likelihood-ratio tests (comparing against


χ2 distribution) following deletion from the model until only those terms that explained significant deviance remained, starting with interactions (e.g., Weber et al. 2011).

Data exploration techniques were applied as described in Zuur et al. (2010) including multipanel scatterplots to look at the pres-ence of collinearity (correlation between explanatory variables). All two-way interac-tions were tested, but results are only pre-sented if found to explain significant vari-ation. Homogeneity and heteroscedasticity were assessed based on a graphical inspec-tion of the residuals (Zuur et al. 2009).

RESULTSWe studied a total of 109 clutches laid over the study period (2008-2009): 35 incubat-ed artificially in eastern beaches (EB), 31 artificially incubated clutches in western beaches (WB) and 43 clutches incubated in situ at western beaches. We could not find 19 of the 62 clutches left in situ, likely due to erosion rainfall and floods. Table 5.1 shows descriptive statistics of number of yolked eggs, hatching success and incubation du-ration of clutches. We used data from 2008 and 2009 only to assess the effects of arti-ficial incubation as no clutches were artifi-cially incubated in WB in 2010.

Table 5.1. Mean (± SD) number of yolked eggs per clutch (YE), hatching success (HS %) and incubation duration (ID days) per clutch of leatherback turtles incubated under artificial (eastern and western beaches) and natural (in situ) conditions in Jaragua National Park during the four years of study. Range in parentheses, n = sample size.

Eastern beaches western beaches In situ

Year YE HS ID FEMALE% YE HS ID FEMALE

% YE HS ID FEMALE %

200865.1±18.4(43-104)

n = 12

53.1±16.4(34.3-79.1)

n = 12

72.6±6.1(66-79)

n = 5

2.5±3.8 (0-8)n = 5

76.3±15.1(47-101) n = 22

42.9±23.9(4.8-85.5)

n = 22

62.7±2.1(60-67) n = 20

41.7±23.5 (8-75)n = 20

71.1±14.8(47-110) n = 22

85.3±19.9(5.0-98.2)

n = 22

64.9±1.8(61-68)

n = 14

53.6±28.5 (0-94)n = 14

200970.1±16.3(24-107)

n = 42

57.5±26.6(0.0-93.6)

n = 42

65.6±3.8(59-74) n = 30

23.4±28.3 (0-85)n = 30

73.2±11.9(54-92)

n = 12

33.7±17.6(6.0-55.8)

n = 12

60.6±3.1(55-64)

n = 11

57.7±26.6 (21-99)

n = 11

83.6±16.5(56-120)

n = 40

73.6±21.8(12.3-97.9)

n = 40

62.7±2.9(57-70)n = 29

72.9±30.7 (0-100)n = 29

TOTAL68.9±16.8

(24-107) n = 54

56.5±24.7(0.0-93.6)

n = 54

66.4±4.7(59-79) n = 35

21.3±27.6 (0-85)n = 35

75.2±13.2(47-101) n = 34

39.6±22.0(4.8-85.5)

n = 34

62.0±2.6(55-67)

n = 31

47.4±25.4 (8-99)n = 31

79.2±16.9(47-120)

n = 62

77.8±21.7(5.0-98.2)

n = 62

63.4±2.8(57-70) n = 43

66.3±31.0 (0-100)n = 43


Effects of incubation method on hatching successIncubation method significantly influenced hatching success (GLMM: χ2 = 76.7, p < 0.001; Figure 5.2a). In situ clutches had greater hatching success than those artificial-ly incubated in WB and in EB (Table 5.1 and Figure 5.2a). We found clutch size (GLMM: χ2 = 12.7, p < 0.05), laying date (GLMM: χ2 = 18, p < 0.01) and incubation duration (GLMM: χ2 = 20.1, p < 0.01) exerted statis-tically significant effects on hatching success. We found incubation method to have an effect on the number of early stage embryo deaths (GLMM: χ2 = 16.1, p < 0.05; Figure 5.2b). Clutches artificially incubated in WB had more early stage dead embryos than in situ clutches and clutches artificially incu-bated in EB (Figure 5.2b). The number of late stage dead embryos in a clutch was also significantly affected by incubation method (GLMM: χ2 = 67.4, p < 0.001; Figure 5.2c). Artificially incubated clutches in EB and WB had more late stage dead embryos than in in situ clutches (Figure 5.2c).

Effect of incubation method on incubation durationAcross the two years, laying dates ranged from 09/03/2008 to 07/06/2008 and 25/03/2009 to 19/06/2009. Laying date was significantly negatively correlated with incubation duration in 2008 with longer incubation durations registered at the be-ginning of the season; however no correla-tion was found in 2009. Incubation method significantly influenced incubation duration (GLMM: χ2 = 16.0, p < 0.01). Clutches ar-tificially incubated in EB had longer incuba-tion duration than in situ clutches and clutch-es artificially incubated in WB (Figure 5.2d).

Figure 5.2. Effect of incubation method (artificially incubated eastern beaches (EB), western beaches (WB), and in situ ) on a) hatching success.clutch-1.year-1, b) number of dead early stage embryos, c) number of dead late stage embryos, and d) incubation duration (days). Shown are fitted values of the model (Mean + SEM), which account for the effect of all significant predictor variables in the model and reflect the true relationship.


Mean incubation temperatures during the whole incubation duration were significant-ly lower in artificially clutches incubated at EB when compared to artificially incubated clutches at WB (Welch’s t-test, t22.9 = - 5.66, P < 0.001; Figure 5.3a). Artificially incubat-ed clutches at WB underwent large cycles of diel temperature variation and in some cases exceed the thermal tolerance range for embryo development under laboratory conditions (35ºC, Ackerman 1997) (Figure 5.3b). Incubation duration was correlated with mean temperature during the ther-mosensitive period (F1, 42 = 15.9, P < 0.001; Figure 5.4) thus it can be used to estimate sex ratio (Broderick et al. 2000).

Figure 5.3. a) Temperatures for the whole incubation period (Mean + SEM) recorded in artificially incubated clutches at the two sites (EB: eastern beaches, and WB: western beaches) in 2008 and 2009, and b) hourly recorded incubation temperature profile of one clutch laid on 15 May 2008 and artificially incubated in WB that reached a maximum temperature of 35.3ºC and had 4.8% of hatching success. Solid line indicates 35ºC; the upper limit of tolerance range for marine turtle embryo development in situ (Ackerman, 1997). Number above bars indicates sample size.

Figure 5.4. Relationship between incubation duration (ID) and mean temperature during the middle third of the incubation duration for 11 artificially incubated clutches in EB (solid circles) and for 33 artificially incubated clutches in WB (open circles). Equation: y = - 3.06x + 152, r2 = 0.28.

DISSCUSION

Effects of artificial incubation on hatching successThe results of our analyses revealed that artificial incubation carried out to protect leatherback turtles clutches in Jaragua Na-tional Park caused changes both in hatching success and sex ratio, when compared to in situ clutches. Artificially incubated clutches, especially those incubated in boxes at west-ern beaches (WB), exhibited lower hatching success than in situ incubated clutches. It is possible that hatching success in some arti-ficially incubated clutches were affected by physical handling of eggs during relocation and transportation (Chan and Liew 1996), particularly in artificially incubated clutches of WB because of it was necessary to trans-port clutches further. However, we believe that the reduced hatching success in artifi-cially incubated clutches was likely due to incubation conditions. Restricted exchange of heat, O2 and CO2 in boxes increas-es embryonic mortality, these differences


in gas exchange might explain differences observed in incubation results (Ackerman 1980; Garret et al. 2010). Most eggs that failed to hatch in in situ clutches contained early stage embryos, mirroring trends seen in other leatherback populations (Eckert and Eckert 1990; Bell et al. 2003) whereas in artificially incubated clutches most eggs contained late stage embryos and many of them appeared to have died just prior to hatching, especially in WB. Metabolic heat generated by embryo activity (Wallace et al. 2004) could have caused this high late em-bryo mortality in WB because several boxes reached extremely incubation temperatures (above 35ºC), which can reduce hatching success (Santidrián Tomillo et al. 2009).

Although the function of yolkless eggs re-mains unclear (Eckert et al. 2012), recent studies have demonstrated that they could be ‘production over-runs’ of oviducts pro-ducing copious albumen for egg clutches and do not have a function related to facil-itate gas exchange or temperature control within the clutch environment (Wallace et al. 2006 and 2007). However, Caut et al. (2006) showed that yolkless eggs had a posi-tive effect on the clutch protecting yolk eggs from insect predation and Patiño-Martinez et al. (2010) propose that space released by dehydration of yolkless eggs may favour ne-onate emergence. Artificial incubation in boxes protects from predation and we kept hatchlings on the sand surface inside the box until they enter into frenzy before being released. For these reasons, and since oth-er translocation programs have not found a significant difference in hatching success between clutches incubated with or without yolkless eggs (Dutton and McDonald 1995; Dal Pont Morisso and Krause 2004), we be-

lieve that not including yolkless eggs in arti-ficially incubated clutches did not markedly affect hatching success.

Effects of artificial incubation on incubation durationAs well as impacting hatching success, the artificial incubation treatments also had an influence upon incubation temperature. The importance of incubation temperature is that it determines incubation duration and hatchling sex (Morreale et al. 1982). The longer incubation durations recorded in EB clutches could be due to environmen-tal conditions, since the rangers’ barrack where the clutches were incubated was lo-cated in a place with higher humidity and likely lower temperatures than in the beach. These longer incubation durations resulted in lower percentage of females produced in EB. However, since no clutches were in-cubated in situ in EB we could not find if there were significant differences on likely sex ratios produced by artificially incubated versus in situ clutches. Taking into account that sand color and quality can have an effect into thermal conditions on beach-es (Naro-Maciel et al. 1999; Weber et al. 2011) it is possible that there are differences in natural sex ratio between beaches. The incubation site at WB is well away from the beach in an area of arid scrub and clutch-es reached high temperatures, resulting in similar incubation durations and sex ratios than in situ clutches.

It is important to note that eastern beach-es are subject to ~100% illegal take of all clutches not relocated (Revuelta et al. 2012) and the shortage of human and economic resources prevents in situ clutch protection here. Therefore, clutch relocation is cur-


rently the only viable option for increasing recruitment, which has been recommend-ed when illegal take of clutches approaches 100% (Mortimer 1999). We improved the previous program leaving the clutches incu-bating in the barracks of each beach and reducing transportation, replacing Styro-foam boxes by polyurethane ones and since 2011, we have been incubating 50% of EB clutches in a hatchery on the beach in or-der to mimic natural sex ratios as closely as possible.

For the artificially incubated clutches in WB, sex ratio proportions were similar to in in situ clutches but had lower hatching suc-cess. Western beaches are less frequented by walkers and rangers’ facilities were installed there during the project, allowing us to re-inforce protection and gradually reduce the number of clutches artificially incubated along the study period. In 2010, consider-ing the low predation rate of in situ clutches we left all clutches incubating on the beach.

The present study fills the gap of leather-back turtle reproductive data in the Domin-ican Republic (Dow and Eckert 2011). In the eastern beaches of the Jaragua Nation-al Park the overall production of hatchlings incubated in boxes during the study period was much greater than would have been expected without protection efforts, but this method probably altered natural sex ratios. Hence, further research is needed to improve conservation measures at these beaches, including enforcement of beach protection and the development of an ef-fective hatchery. Our results also indicated that artificial incubation at WB of JNP was an ineffective conservation strategy accord-ing to the low levels of hatching success re-corded; thus current conservation is based

in in situ clutch protection. However, WB are subject to unregulated tourism activity and development and are under the threat of the expansion of bauxite and limestone mining (Wielgus et al. 2010). Hence, this conservation strategy will be only possible with an institutional control over the influx of tourism and coastal development.

It is possible that, in the future, relocation and artificial incubation will be an option to avoid the extirpation of threatened ma-rine turtle nesting populations due to hab-itat loss by coastal development or to the predicted rise of sea level. Nonetheless, our findings highlight the importance of monitoring and assessing mortality and sex ratio in conservation programs using these methods.

ACkNOWLEDGEMENTSThe present study formed part of a con-servation project funded by: the Spanish International Cooperation Agency (AECI, projects: A/2991/05 and A/5641/06), the Spanish Ministry of Education and Scienc-es (CGL2006-02936-BOS), the General Foundation of the University of Valencia, and the European Union (Marie Curie grants, FP6 and 7). We especially thank the Dominican NGO Grupo Jaragua and The Nature Conservancy (TNC). The project was supported by the Dominican authori-ties (Office of Protected Areas and Biodi-versity of the Environmental Ministry) and the rangers of the National Parks. All works and manipulation of eggs and hatchling have been made under the permits of Envi-ronmental Ministry. We are very grateful to fieldwork assistants especially to “Blanco”


Turbí and the other rangers for their help during fieldwork in Jaragua and their hard work in the conservation of marine turtles. We extend our thanks to the Dominican and international students, volunteers and local people who helped in many ways. We also thank D. Weber and D. Kross for many helpful comments on the statistics. J. Tomás and J.A. Raga are supported by project Prometeo/2011/40 of Conselleria de Educacio (Generalitat Valenciana) and project CGL2011-30413 of the Spanish Ministry of Sciences and Innovation. The manuscript was improved as a result of the input of the Editor and two anonymous re-viewers.


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6. CHAPTER VI: RUNNING AGAINST TIME: CONSERVATION OF THE REMAINING HAWkSBILL TURTLE (ERETMOCHELYS IMBRICATA) NESTING POPULATION IN THE DOMINICAN REPUBLIC

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ChApTEr VI: RUNNING AGAINST TIME: CONSERVATION OF THE REMAINING HAWkSBILL TURTLE (ERETMOCHELYS IMBRICATA) NESTING POPULATION IN THE DOMINICAN REPUBLIC

CHAPTER VI:

RUNNING AGAINST TIME: CONSERVATION OF THE REMAINING HAWkSBILL TURTLE (ERETMOCHELYS IMBRICATA ) NESTING POPULATION IN THE DOMINICAN REPUBLIC

Ohiana Revuelta1, Yolanda M. León2, 3, Francisco J. Aznar1, Juan A. Raga1 and Jesús Tomás1

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Published in Journal of the Marine Biological Association of the United Kingdom (2013) volume 93: 1133-1140





06

Chapter VI: Running against time: conservation of the remaining hawksbill turtle (Eretmochelys imbricata) nesting population in the Dominican Republic I 133

ABSTRACTSaona Island hosts the last hawksbill turtle (Eretmochelys imbricata) nesting population in the Dominican Republic, which has experienced a severe decline in the last decades, mostly due to illegal egg take. Here we present the results of an artificial incubation pro-gramme started in 2007 to protect the clutches from human predation. A preliminary survey in 2006 showed that about 60% of clutches laid were taken by humans. Over the study period (2007-2010) we recorded 400 clutches, of which 38.2% were predated by humans, 40.7% were artificially incubated and 21% were incubated in situ. Overall, the artificial incubation programme allowed the release of 12,340 hatchlings. No differences were found in hatching and emergence success between clutches incubated in situ and clutches artificially incubated. However, incubation temperatures and incubation dura-tions recorded suggest a male-biased hatchling sex-ratio in artificially incubated clutches. Although artificial incubation may mitigate the effect of egg take, our results indicate that other measures, such as clutch relocation to protected sections of the beach should be tak-en. Beach patrolling and education are currently implemented so that artificial incubation will be eventually phased out in favour of in situ incubation.


INTRODUCTIONThe hawksbill marine turtle, Eretmochelys imbricata, is listed under the IUCN Global Red List as critically endangered (IUCN 2011). In the Caribbean, this species has been exploited historically for its meat, eggs and shell (Meylan 1999; Carrillo et al. 1999; Fleming 2001) leading to a notice-able reduction in nesting populations in the region (Meylan 2002; McClenachan et al. 2006). However, the regional management unit of Caribbean hawksbill is currently characterized as low risk but as high threat (Wallace et al. 2010), and some populations are thought to be increasing (Puerto Rico: Meylan 1999; Antigua: Richardson et al. 2006; Barbados: Beggs et al. 2007; Guade-loupe: Kamel and Delcroix 2009). Howev-er, rookeries without protection in this basin are seriously threatened, particularly due to egg take (Lagueux and Campbell 2005).

Increasing hatchling production is a neces-sary component of any strategy to recover depleted marine turtle populations (Dut-ton et al. 2005; Sarti Martínez et al. 2007). When clutches are at risk, relocation to dif-ferent sections of the beach or to protected hatcheries have been common management tools in many marine turtle conservation programmes (Pritchard 1995; Kornaraki et al. 2006; Marcovaldi et al. 2007). However, this is not a perfect solution, as it has been reported that relocated clutches may expe-rience lower hatching success and altered sex-ratios (Godfrey and Mrosovsky 1999; Mortimer 1999; Pintus et al. 2009). Never-theless, the intense pressures from human egg take in undeveloped countries leaves them with no other option than relocating clutches and using hatcheries (García et al.

2003; Liles et al. 2011) or even, in extreme situations, to incubate clutches in boxes (Dutton et al. 1985). Therefore, clutch con-servation strategies should rely on the eval-uation of local characteristics, and should be adapted to the specific scenarios where they are intended to be applied (Eckert 1999; Kornaraki et al. 2006).

In the Dominican Republic (DR, North Caribbean), current estimates suggest a se-vere decline of the hawksbill nesting pop-ulation, with sporadic nesting in few places around the country and only one remain-ing stable stock on the south coast of Sa-ona Island, south-east DR (with a mean of 23 females nesting per year: Revuelta et al. 2012). According to the estimation of 5000 females nesting annually in the Caribbean (Meylan 1999), Saona Island hawksbill stock would represent close to 0.5% of the total number of nesting fe-males in the region. The greatest threat to this nesting rookery comes from illegal egg take and deliberate capture of adults for meat (Revuelta et al. 2012). Preliminary surveys in 2006 revealed that up to 60% of 41 nests recorded on the island were taken by humans, with up to 100% (n = 13) taken near the principal human settle-ment, Mano Juan (Revuelta et al. unpub-lished data). This occurs even though it is illegal to harvest turtles or their products in the DR since 1966, and that Saona is part of a protected area since 1975 (Del Este National Park). This situation could potentially impact rookeries on a region-al scale since Dominican hawksbills seem to disperse to several feeding grounds throughout the Caribbean after nesting; including distant areas in waters of Nica-ragua and Honduras (Hawkes et al. 2012).


It is well known that artificial incubation in boxes may introduce masculinizing bi-ases in resulting hatchlings (Morreale et al. 1982; Dutton et al. 1985; Whitmore and Dutton 1985). This is because sex in marine turtles is determined by the temperature at which eggs are incubated during the mid-dle third of incubation, which is referred to as the thermosensitive period (Yntema and Mrosovsky 1982). When boxes are used, the temperature during this period is often lower than pivotal temperature (that which produces 1:1 sex ratio: Mrosovsky and Pieau 1991), thus resulting in a greater proportion of male hatchlings (Mrosovsky 1994). However, given the high predation levels observed, in 2007, we initiated a pro-gramme of artificial incubation in boxes as an urgent measure to protect the clutches and increase hatchling recruitment while implementing other conservation measures, such as beach patrolling and education.

Here we present the results of a four year monitoring and conservation programme of the hawksbill nesting population of Saona Island, including the artificial in-cubation programme. Our main research questions were twofold: (1) are there differ-ences in hatching and emergence success between artificially incubated clutches and in situ incubated clutches?; and (2) what is the sex-ratio produced in the artificially in-cubated clutches?


Study areaThe present study was conducted from 2007 to 2010 in Saona Island (18º07’N 68º44’W; Figure 6.1), which is included in the Del Este National Park (south-east DR). With an area of 110 km2, Saona is the largest island adjacent to the DR. There is one permanent human settlement in Sao-na, Mano Juan village, with a population of 300 inhabitants (Figure 6.1). Hawksbill nesting occurs across all of the 26.6 km of the south coast sandy beaches of Saona Is-land. However, nesting activity is concen-trated in 12 narrow fine-grained coralline, white sand beaches (altogether comprising a total of 15 km; Table 6.1) interspersed with rocky areas. The neritic sea adjacent to the nesting beach is composed of coral reefs and seagrass beds. For the most part, beach vegetation is dominated by coconut (Cocos nucifera) plantations, sea purslane (Se-suvium portulacastrum), sea rosemary (Suriana maritima), sea grape (Coccoloba uvifera), goat’s foot creeper (Ipomoea pes-caprae) and native grasses.


Figure 6.1. Hawksbill turtle nesting areas (n = 5) and beaches surveyed (n = 12) in Saona Island in Del Este National Park, Dominican Republic. Inserted in the map is shown the location of the study area within the Caribbean Sea.

Beaches were grouped into 5 sampling ar-eas: (1) Del Toro, with 1530 m length and mean beach width (range) of 7.9 (4-18) m; (2) Mano Juan, 2990 × 8 (2-20) m; (3) Boca Chica, with 12.9 × 4.4 m; (4) Canto de la Playa area, 4263 × 18 (8-30) m; and (5) Faro Punta Cana, 6084 × 10 (3-20) m (Figure 6.1). We patrolled the beaches at least once per week throughout the year, but in 2008 this was increased to 3-4 sur-veys per week during the peak nesting pe-riod (June to November). Also, in 2009, the farthest area (Faro Punta Cana; see Figure 6.1) was visited only monthly due to logistic limitations.

We patrolled the beaches, by foot to de-tect all recent tracks of nesting females. All emergences of turtle females to nest were

recorded. Clutches were confirmed by the presence of eggs in the nest chamber, or by signs of egg predation at the nesting site (see below). For each nest we recorded its GPS location, and position on the beach (open sand, vegetation border and within vege-tation). We measured minimum distances from the nest to the high tide line. We clas-sified nests as intact (eggs in nest chamber), or predated (no eggs in nest chamber, pres-ence of probing sticks, eggshells and human footprints littering the nesting site). Owing to high levels of human egg take, the ma-jority of clutches detected (n = 163) were removed for artificial incubation. In order to investigate the potential effects of artifi-cial incubation on hatching and emergence success we studied 49 clutches incubated in


situ, using them as controls. To reduce pre-dation, all nests left in situ were camouflaged by erasing turtle tracks and other signs of nesting activity.

Clutches moved for artificial incubation were carefully excavated by hand, and the eggs were relocated into polyethylene exte-rior boxes with polyurethane foam filling. Two to three centimetres of beach sand was put at the bottom and sides of the box to prevent contact of eggs with walls. We placed a TinyTag temperature logger (Gemini Data Loggers UK Ltd., Model TGP-4017; accuracy +0.38C) inside a sub-sample of 22 artificially incubated clutches. Loggers were programmed to record tem-peratures at sampling intervals of 1 hour throughout the incubation duration. Extra care was taken when moving clutches, lim-iting shake and vibrations to avoid dislodg-ing the embryo (Chan 1989; Almeida and Mendes 2007). Each box was labelled with a code indicating beach name, the laying date and number of eggs incubated. Boxes were placed in a facility located about 50 m from the beach of Mano Juan village. This facility (4 × 3 m (length × width)) had a sand-beach floor, walls made of wire mesh and corrugated metal panels, and roof made from palm leaves to protect boxes against rain and flooding. The boxes were checked daily throughout the incubation duration and lids opened for two to three hours a day to allow air to circulate in the otherwise air-tight box. Hatchlings were re-leased at sunset one day after they emerged when they enter into frenzy.

To detect effects on hatchling fitness that might be due to artificial incubation, we compared the size and weight of hatchlings produced in boxes with those of hatchlings

from other nesting rookeries in the Carib-bean (e.g. Mona island, Puerto Rico: van Dam and Diez 1998; Cuba: Moncada et al. 1999; British Virgin Islands: Hillis and Phil-lips 1996). Random samples of 20 hatch-lings each from 24 artificially incubated clutches in 2008 and 8 artificially incubated clutches in 2009 were measured (straight carapace length (SCL)) to the nearest 0.1 cm with a calliper, and weighed to the near-est 0.1 g with an electronic scale.

Artificially incubated and in situ clutches were studied, and clutch size, hatching and emergence success were considered accord-ing to definitions in Miller (1999). In this study, it could not be known in advance whether the clutches we collected for arti-ficial incubation were a random subset of all the clutches on the beach. A way to shed light on this question could be to check whether clutch size differed between arti-ficially incubated and in situ clutches (note that, in the latter, clutch size could only be determined after emergence of hatchlings). A two-way analysis of variance (ANOVA) revealed that the mean number of eggs per clutch did not differ with incubation type (F(1,157) = 0.85, P = 0.357) or among years (F(3,157) = 2.46, P = 0.06).

We tested whether incubation type affect-ed hatching success or emergence suc-cess. Since data were gathered in 4 years, we used a full factorial two-way ANOVA, with ‘incubation type’ (in situ, artificial) and ‘year’ as fixed and random factors, respectively. Before statistical testing, data about hatching and emergence success were arcsin-transformed to achieve nor-mality and homoscedasticity (Sokal and Rohlf 1995).


For each clutch, incubation duration was defined as the number of days between egg-laying and the first emergence record-ed. We used temperature data to provide an estimation of sex-ratio produced in the artificially incubated clutches. Since pivot-al temperature has not been determined for hawksbill turtle in Saona, we used the published curve relating incubation tem-perature and hatchling sex-ratio derived

from laboratory incubated eggs from the closest hawksbill nesting area at Mona Is-land, Puerto Rico (Mrosovsky et al. 2009). We read off the means of temperature re-corded in the thermosensitive period of the 22 clutches on the Mona laboratory curve to estimate female proportions. Examples of the use of this method and its validation are given in Wibbels (2003) and Öz et al. (2004).

RESULTSTable 6.1. Number of nests and distribution of hawksbill turtle nesting density per beach in Saona Island, during 2007-2010. BL, beach length (km); MANR, mean annual nests rate (nests.years-1); ND, nesting density (nests.Years-1.kms-1).

Area Beach BL (km) N MANR ND

Toro Toro-Subidero 1.5 39 9.0 6.0

Mano Juan Farito-Valdés 0.7 65 16.2 23.1

Cementerio 0.29 52 13.0 44.8

Mano Juan 2.0 18 4.5 2.2

Boca Chica Boca Chica 0.01 23 5.7 438.4

Canto Playa Canto Playa 1.2 66 16.5 13.7

Griegos 3.1 11 2.7 0.8

F.Punta Cana M. Careyes 0.8 6 1.5 1.8

Cuerno 1.5 22 5.5 3.7

Caletón Sucio 0.5 20 5.0 10.0

F.Punta Cana 2.1 68 17.0 8.1

Jayuya 1.2 10 2.5 2.1


Nesting density and nest site selection Hawksbill turtles laid a mean of 100 nests.year-1 (± standard deviation (SD) 8.4: range 93 to 111) on Saona Island. The highest mean annual nesting rate occurred at Mano Juan area (33.7 nests.year-1) followed by Faro Punta Cana area (31.5 nests.year-1), Canto de la Playa area (19.2 nests.year-1) and Del Toro area (9.0 nest.year-1). The mean nest-ing linear density over the 15 km of avail-able nesting habitat was 6.6 nests.year-1.km-1. Apart from Boca Chica, an isolated small beach surrounded of rocky areas, with an average of 5.7 nests.years-1, the highest nesting density reported was on Mano Juan area (11.3 nests.year-1.km-1), followed by Del Toro area (6 nests.year-1.km-1), Faro Punta Cana (5.2 nests.year-1.km-1) and Canto de la Playa (4.5 nests.year-1.km-1) (see Table 6.1). We did not find significant correla-tion between beach width and number of clutches laid (rs = 0.451, P = 0.141, n = 12).

Saona hawksbills laid their eggs mainly in zones with at least some vegetation (within vegetation: 26.5%, n = 75; vegetation bor-der: 56.0%, n = 158; open sand: 17.4%, n = 49). Overall, the mean distance from the nest to the high tide line was 8.9 m (SD = 5.3; range: 0.5-35 m). The beach with high-est mean distance (16.7 m; SD = 7.0; range: 8-30) was Canto de la Playa beach, in Can-to de la Playa area; and the beach with low-est mean distance (5.7 m; SD = 3.2; range: 2-13) was Cementerio beach, located in the Mano Juan area (Figure 6.1).

Predation levels and fate of clutchesIn the four-year period we recorded a total of 400 hawksbill clutches in the 5 sampling areas. Considering the extensive experience detecting hawksbill tracks by the team and the low levels of nesting activity, we believe that weekly surveys allowed us to record close to 100% of clutches laid.

The majority of recorded clutches were artificially incubated (n = 163), while 146 clutches had already been predated by hu-mans when found and 91 were left in situ (Figure 6.2). Humans took an additional 7 of these 91 clutches left in situ, and other 15 clutches of them were affected by tropi-cal storms. We could not find (likely due to erosion and wash-out) another 20 of the 91 clutches left in situ. These events left a total of 49 clutches incubated in situ for the study.

Hence, in total we include in the study 365 clutches of three different fates: (i) predated by humans (n = 153); (ii) incubated artifi-cially (n = 163); and (iii) incubated in situ (n = 49). However, for the estimations of hatching and emergence success we only considered clutches that were studied by the authors (n = 119) and (n = 46) artifi-cially incubated clutches and in situ clutch-es, respectively. Figure 6.3 shows the annual variation on the percentages of these cate-gories. All the clutches that were not artifi-cially incubated or camouflaged in situ were found predated by humans in all beaches of the five areas of study. When humans pre-dated a clutch, 100% of eggs were always taken. No evidence of natural predation (i.e. by crabs, ants, or feral pigs) on eggs was observed during the study period.


Figure 6.2. Clutch fate: number of clutches laid and their fate in Saona Island during the period 2007-2010. n, total number of clutches; predated, number of clutches predated by humans; artificial incubation, number of clutches removed for their incubation in boxes; in situ incubation, number of clutches camouflaged and left incubating on the beach; predated post-camouflage, number of clutches predated by humans after we camouflaged it; in situ studied, clutches found and studied after being camouflaged; affected by hurricanes, number of clutches incubating on beach when a hurricane made landfall in the island; not found, number of camouflaged clutches we could not find again due to beach dynamics.

Figure 6.3. Percentages of clutches that were incubated artificially, clutches that were poached and clutches that hatched successfully in situ from the total of studied clutches every year (n = 84; n = 94; n = 92 and n = 95) respectively.

Find a nest during surveys

n = 400

Artificial incubation

n = 163

in situ incubation

n = 91

in situ studied n = 49

Affected by hurricanes

n = 15

Not foundn = 20

Predated post camuflage

n = 7

Non predated post camuflage

n = 84

Non predated n = 254

Predated n = 146


Hatching success and emergence successHatching and emergence success of 46 in situ incubated clutches and 119 artificially incubated clutches are shown in Table 6.2. We did not detect significant effects of artifi-cial incubation in hatching success between years (two-way ANOVA: F (3,157) =1.45, P = 0.383), incubation type (F(1,157) = 0.7, P = 0.45); or their interaction (F(3,157) = 0.52, P = 0.669). Likewise, there was no significant difference in emergence success between years (two-way ANOVA: F (3,157) = 1.1, P = 0.469), incubation type (F(1,157) = 1.28, P =

0.327); or their interaction (F(3,157) = 0.68, P = 0.563) compared to in situ incubation.

In total, 12,340 hatchlings were produced under artificial conditions and released to the sea (1731 in 2007, 4522 in 2008, 2664 in 2009 and 3423 in 2010). Mean carapace length and weight of hatchlings was 3.8 ± 0.1 cm (range: 3.2-4.2) and 14.8 ± 1.2 g (range: 10.5-18.7) in 2008 (n = 480) and 3.8 ± 0.2 cm (range: 3.0-4.0) and 15.1 ± 1.1 g (range: 12.6-18.2) in 2009 (n = 160).

Table 6.2. Number of yolked eggs (YE), hatching success (HS), emergence success (ES) and incubation duration (ID) per clutch of hawksbill turtles reared under natural (in situ) and artificial conditions in Saona Island during the four years of study. N, number of clutches.

YEAR N YE HS ES N ID

In situ Range Mean SD Range Mean SD Range Mean SD

2007 9 1092 37.0-92.6 72.3 16.6 35.2-92.6 71.1 17.6 4 57- 64 60.6 3.4

2008 9 1392 5.1 - 96.5 73.3 30.9 5.1 - 96.5 72.1 30.3 4 68- 90 80.7 9.3

2009 7 929 52.8-91.4 78.3 12.8 34.9-91.4 72.0 21.0 0 — — —

2010 21 3034 7.1 - 98.3 71.8 25.5 7.1 - 98.3 67.0 25.7 16 50- 79 60.9 8.0

TOTAL 46 6447 5.1 - 98.3 73.2 23.1 5.1 - 98.3 69.5 24.0 24 50- 90 64.2 10.6

Artificial

2007 20 2513 28.9-99.2 72.3 20.3 28.9-99.2 68.9 20.2 14 52- 66 59.8 3.7

2008 41 5667 44.5-98.5 81.8 13.9 32.1-98.5 79.8 14.7 25 53- 85 62.6 9.5

2009 25 3501 40.4-97.6 77.5 13.3 39.4-97.6 76.1 13.9 21 56- 71 59.4 4.1

2010 33 4588 27.6-100 76.7 20.6 27.6 -100 74.6 20.4 24 53- 85 62.3 8.4

TOTAL 119 16269 27.6- 100 77.9 17.1 27.6- 100 75.5 17.6 84 52- 85 61.3 7.3


Incubation temperature and incubation duration Since beaches were not daily patrolled, re-cording the exact laying date was not pos-sible in all cases. However, we were able to determine incubation duration for 84 out of the 119 artificially incubated clutches and 24 out of 46 in situ incubated clutches studied (Table 6.2). Temperatures in artificially incu-bated clutches ranged from 22.2 to 36.4ºC, with mean values ranging from 26.8 to 30.2ºC (n = 22). During the thermosensitive period temperatures recorded ranged from 22.2 to 33.8ºC, with mean values ranging from 25.7 to 29.8ºC. Incubation duration was strongly correlated with mean tempera-ture during the thermosensitive period (F = 70.7, P, 0.001, df= 1; Figure 6.4).

In 21 of the 22 artificially incubated clutch-es mean temperature during the thermo-sensitive period was lower than the pivotal temperature derived from laboratory stud-ies (29.6ºC: Mrosovsky et al. 2009) (Figure 6.5), suggesting a male bias in artificially incubated clutches. The highest estimated female percentages (42% and 73%) were recorded in two clutches laid during the first half of August, and the estimated per-centage of females was 0% in all artificial-ly incubated clutches laid after the second half of August (Figure 6.5). Accordingly, we found a great intra seasonal variation in incubation duration in artificially incubat-ed clutches every year, with longer incuba-tion duration during colder months (Octo-ber-December).

Figure 6.4. Relationship between incubation duration and mean temperature during the thermosensitive period in the 22 artificially incubated clutches controlled during the study (Equation: ID = 211-5.26 mean temperature during the thermosensitive period, r-square = 0.78).


Figure 6.5. Laying date, mean incubation temperature during thermosensitive period (dots) and estimated female proportion (bars) of artificially incubated clutches (n = 22) in which temperature was monitored during the study period.

DISCUSSIONThe present study provides, for the first time, detailed information about the repro-ductive biology of the remaining, threat-ened hawksbill nesting population of Saona Island, as well as the results of the conser-vation programme that we carried out to mitigate egg take.

Nesting density and nest site selectionHawksbill females on Saona mainly nest in vegetated areas, this has also been not-ed in other studies of the species in other Caribbean rookeries (Pérez-Castañeda et al. 2007; Kamel and Delcroix 2009). Oth-er authors have proposed that this could be because areas with vegetation might be less compact making hatchling emergence easier; also, the shade provided by vegeta-

tion might help keep temperature constant (Horrocks and Scott 1991; Kamel and Mrosovsky 2006). Moreover, the two beach-es with the highest nest density in Saona (Boca Chica and Cementerio) have certain characteristics which make them suitable for nesting: an accessible vegetation area, a reef in front of the beach which protects them against surf and makes beach access easier, and fine grain sand that allows tur-tles digging their nests easily.

Predation levels and fate of clutchesOur results show the need to continue pa-trol and conservation work in Saona, since most in situ clutches not promptly camou-flaged or relocated for artificial incubation, were predated by humans. In fact, in 2009, as a result of the decrease in surveys at the Faro Punta Cana area, 28 out of the 30


clutches laid were taken. In other words, the current level of egg take is unsustain-able for the long term preservation of this nesting population.

Nesting beaches in Saona are scattered along the coastline and some of them can only be reached by boat, considerably rais-ing monitoring costs. Also, the funding available did not allow the employment of personnel to make daily patrols over all the nesting beaches or nest surveillance. This left us with two options, i.e. relo-cating or camouflaging all clutches that were non-predated at the time of finding. Camouflage was considered to be an in-sufficient conservation measure since sub-stantial amounts of camouflaged clutches were eventually predated by humans. The same situation has been described in oth-er studies with similar conditions (Sato and Madriasau 1991; Lagueux et al. 2003). On the other hand, tropical storms that hit the island affected camouflaged clutches (4 af-fected by Noel in 2007, 7 by Hanna in 2008 and 4 by Earl and Fiona in 2010) making excavation and study impossible. It has been reported that excessive rainfall and floods can greatly reduce egg viability, thus reducing clutch survivorship (Martin 1996; Kamel and Delcroix 2009). Obviously, these threats are natural and unavoidable, but they add significant losses to the already depleted population. For all these reasons, and due to shortage of human and eco-nomic resources, we decided to incubate clutches in polyurethane hermetic boxes with constant monitoring and protection, as an alternative option to ensure hatchling recruitment.

Effects of artificial incubationPhysical handling of eggs can reduce hatching and emergence success in relo-cated clutches (Eckert and Eckert 1990; Pintus et al. 2009). However, in the present study, we found no significant differences in hatching and emergence success between in situ and artificially incubated clutches, indi-cating that handling was correctly carried out and that incubation conditions in boxes was suitable for hawksbill embryos develop-ment, as reported also for other species (e.g. Whitmore and Dutton 1985).

Concerning hatchling fitness, our results suggest that artificially incubated hatch-lings from Saona are smaller than hatch-lings from other Caribbean natural nesting rookeries, Mona Island (SCL = 4.0 cm: van Dam and Diez 1998), Cuba (SCL = 4.0 cm: Moncada et al. 1999) and the Virgin Islands (SCL = 4.1 cm: Hillis and Phil-lips 1996). However, the weight of Saona hatchlings seems to be similar to those from other Caribbean rookeries (14.8 g, Mona Island and 14.7 g, Virgin Islands).

Nonetheless, the use of boxes for incuba-tion entails the risk of bias in the sex-ratio in favour of males (Dutton et al. 1985; Whit-more and Dutton 1985). Mean tempera-tures of 21 artificially incubated clutches during the thermosensitive period were be-low the pivotal temperature reference value from Mona Island (29.6ºC), suggesting that those temperatures are likely to have result-ed in a male-skewed clutch sex-ratio. More-over, since we found a clear relationship between temperature and incubation dura-tion recorded in Saona, the long incubation durations recorded are also indicative of a male-skewed sex-ratio. The low percentag-


es of female hatchling production inferred from our clutch temperature data, calls for urgent changes in the conservation strategy adopted in Saona, since a lengthy lack of female production would prevent popula-tion growth.

Forward strategiesDue to the rampant egg take, in situ incu-bation on the beaches is not an acceptable management practice for a long term con-servation of this population. Hence, the primary goals in the near future would be: (1) to increase female production in boxes; and (2) to start a programme of relocation of clutches to safer areas (Pritchard 1995; DeGregorio and Southwood Williard 2011) while progressively reducing incubation in boxes. For the first goal, boxes are being ex-posed to direct sunlight during a few hours a day to raise incubation temperatures and increase the proportion of females pro-duced (see Chantrapornsyl 1992). For the second goal, financial resources must be invested to strengthen the protection of a beach section near Mano Juan village as a safe area for clutch relocation.

Further research on beach temperature pat-terns is needed to increase understanding of the natural sex-ratio at Saona in order to improve conservation measures. Togeth-er with the artificial incubation programme we carried out additional measures such as hiring people from the local community to work as field technicians, organizing work-shops and meetings with authorities, as well as environmental education and awareness programmes. These measures have appar-ently positively influenced local attitudes and perceptions toward marine turtles in Saona (White et al. 2011) which we strongly

believe will favour a future programme of clutch relocation on the beach.

ACkNOWLEDGEMENTSThe present study is included as part of a conservation programme funded by the Spanish International Cooperation Agency (AECI, projects: A/2991/05 and A/5641/06), the Spanish Ministry of Ed-ucation and Sciences (CGL2006-02936-BOS), the General Foundation of the University of Valencia, and the European Union (Marie Curie grants, FP6 and 7). We specially thank the support of the Domini-can non-governmental organization Grupo Jaragua and of The Nature Conservancy (TNC). The project counts on the support of the Dominican Environmental author-ities (Office of Protected Areas and Biodi-versity of the Environmental Ministry) and the rangers of the Del Este National Park. We are very grateful to fieldwork assistants especially to Pelagio Paulino ‘Negro’ and Bernardo Paulino ‘Mico’ for their help during fieldwork in Saona and their hard work in the conservation of Saona marine turtles. We extend our thanks to the Domin-ican and international students, volunteers and local villagers who helped in different tasks of the project and who lent their help and friendship during the work in the Do-minican Republic. J. Azar, J.A. Raga and J. Tomás are supported by project Prometeo/ 2011/40 of Conselleria de Educacio (Gen-eralitat Valenciana) and project CGL2011-30413 of the Spanish Ministry of Sciences and Innovation. We thank an anonymous referee for helpful comments and sugges-tions that improved the manuscript.


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7. CHAPTER VII: THE VALUE OF ENDANGERED SPECIES IN PROTECTED AREAS AT RISk: THE CASE OF THE LEATHERBACk TURTLE IN THE DOMINICAN REPUBLIC

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ChApTEr VII: THE VALUE OF ENDANGERED SPECIES IN PROTECTED AREAS AT RISk: THE CASE OF THE LEATHERBACk TURTLE IN THE DOMINICAN REPUBLIC

CHAPTER VII:

THE VALUE OF ENDANGERED SPECIES IN PROTECTED AREAS AT RISk: THE CASE OF THE LEATHERBACk TURTLE IN THE DOMINICAN REPUBLIC.

Ohiana Revuelta1, Yolanda M. León2,3, Juan A. Balbuena1, Annette C. Broderick4, Pablo Feliz3, Brendan J. Godley4, Juan A. Raga1 and Jesús Tomás1, 4

__

Published in Biodiversity and Conservation (2014) doi: 10.1007/s10531-014-0682-x.

__1 Cavanilles Institute of Biodiversity and Evolutionary Biology, University of Valencia, P.O. Box

22085, E-46071 Valencia, Spain.2 Instituto Tecnológico de Santo Domingo, Urb. Galá, Santo Domingo, Dominican Republic.3 Grupo Jaragua, El Vergel, Santo Domingo, Dominican Republic.4 Centre for Ecology and Conservation, School of Biosciences, University of Exeter Cornwall Campus,

Penryn, UK.

07

Chapter VII: The value of endangered species in protected areas at risk: the case of the leatherback turtle in the Dominican Republic I 155

ABSTRACTProtected areas are considered essential elements for global biodiversity conservation. They may not necessarily result in an effective conservation of resources in developing countries due to lack of funding for management and enforcement. In addition, poor governance aligned with conflicts of economic interests related to their use can further threaten their integrity and persistence. In the Dominican Republic, the western beaches of the Jaragua National Park, a protected area which is also part of a UNESCO Biosphere Reserve, have been proposed for development using a mass-tourism model. One of the most charismatic species found in this area is the leatherback turtle (Dermochelys coriacea). In the present study, we assess hatching success, and factors affecting it, to determine the reproductive value across the area for the leatherback turtle. The main factors found driving hatching success at the study beaches are beach sector, incubation duration, date of lay and clutch size. Our results show that clutches in La Cueva (located in the buffer zone of the park) and Bahía de las Águilas (located inside the limits of the park) have an unusually high hatching success (~75%) for this species, highlighting the importance of increasing protection efforts at these sites. We strongly recommend including La Cueva inside the limits of the Jaragua National Park.


INTRODUCTIONProtected areas remain a cornerstone of global conservation efforts to preserve di-minishing wildlife species and their habitats (CBD 2010; Butchart et al. 2012). In the Caribbean, the establishment of Marine Protected Areas (MPAs) to protect natural resources, unique habitats and threatened species have proliferated over the last few decades (Guarderas et al. 2008). Neverthe-less, in many countries not all protected ar-eas have management plans, and when they exist, the national authorities responsible for their protection are often under-resourced, making protection less effective (e.g. Ervin 2003; Buitrago et al. 2008).

The Dominican Republic (DR) is a develop-ing country experiencing a rapid increase in international tourism, particularly focussed on coastal areas of the north and southeast regions (León 2007; Wielgus et al. 2010). In recent years, this mass-tourism model and the expansion of existing bauxite and limestone mining have been proposed as valuable ways to enhance the impoverished economy of the southwest of the country (Wielgus et al. 2010). It is known that these activities could result in an adverse impact on natural environments such as pollution and sand mining, with detrimental effects on reefs and seagrass ecosystems (Geraldes 2003; Grandoit 2005).

At the southwest of the country, the Bahía de las Águilas hosts one of the last refuges of coastal marine fauna and flora available to the Dominican Republic. The ecosystems are thought unlikely to be able to support mass tourism (Wielgus et al. 2010). This area has unique vegetation with several

endemic plants and, among its marine eco-systems; there are the most extensive and best preserved seagrass beds in the coun-try. Coral reefs are found a short distance from the coast, so it is highly probable that these reefs would be greatly affected by any land-based pollution, particularly resultant from mining activity. Finally, the bay hosts the reproductive areas of rhinoceros iguana (Cyclura cornuta) and marine turtles (Rupp et al. 2005; Revuelta et al. 2012).

One of the most charismatic species in this area is the leatherback turtle (Dermochelys co-riacea). Although populations of this species seem to be recovering in the North Atlan-tic, and it has assigned to be under “low risk a low threat” (Wallace et al. 2011), in some rookeries this species face serious threats, such as egg take that compromise their con-servation (e.g., Patiño-Martínez et al. 2008). Moreover, the leatherback turtle exhibits what is considered a very low hatch success rate (~50%) in comparison to other marine turtle species (Bell et al. 2003; references in Eckert et al. 2012).

In the past, the leatherback turtle was re-ported nesting widely around the coast of the DR (Ottenwalder 1981), but today the last important nesting areas for the species are located in the beaches of the Jaragua National Park (JNP; Revuelta et al. 2012). There are two leatherback nesting areas in the Park, the beaches adjacent to the Ovie-do lagoon at the east, which experienced 100% egg take by local people, and Bahía de las Águilas and La Cueva beaches in the west of the Park (Revuelta et al. 2012), but La Cueva is only currently within the buff-er zone of the Park (see study site section). Since 2006, a cooperative marine turtle conservation project has been carried out


at the JNP and, as a result, leatherback clutches are currently monitored for pro-tection by rangers at the western nesting beaches. However, the strong pressure to bring mass tourism to these beaches in the form of high capacity resorts represents a potential threat to protected fauna of the Park (Rupp et al. 2005) including nesting marine turtles. The JNP was created in 1983 and since 2002 it has been part of the UNESCO Jaragua-Bahoruco-Enriquillo Biosphere Reserve. Multiple national laws as well as International agreements rati-fy protection for the Bahía de las Águilas territory and/or its natural resources: The General Environmental Law 64-00, the Sectorial Law for Protected Areas 202-04, The Convention on Biodiversity, and The Cartagena Protocol for Wildlife Areas with Special Protection. In Bahía de las Águi-las, building and extraction of animals or plants are prohibited, but visits of tourists by boat or by 4x4 vehicles into the beach are permitted. A ranger post is placed at the entrance of the bay, and rangers patrol the beach regularly. In the buffer zone of the park there is no patrolling and access to the beach is uncontrolled.

Here we set out (1) to study hatching suc-cess and factors affecting it to determine the reproductive value across the area for the leatherback turtle and, (2) to ascertain the importance of both nesting beach-es for this threatened species to reinforce conservation in the face of anthropogenic threats at site.


Study siteThe present study was conducted between 2007 and 2010 at the western beaches of The Jaragua National Park (JNP, N17° 57’, W71° 39’) a protected area of 1374 km2 (of which 905 km2 is marine reserve) situated in the south-western corner of the DR (Figure 7.1). Bahía de las Águilas is a 4.4 km sandy bay located inside the park but is also considered as a national recreational area according to local laws and is frequented by several thousand Dominican and international tourists per annum. La Cueva beach (2.5 km in length) is located in the northern buffer zone of the park. A small village (around 20 inhabitants) and a rangers’ post are located between these two beaches that are separated by a rocky zone of 1 km length.


Figure 7.1. a) Map of the Jaragua National Park (SW DR) showing the park limits and the areas of recreational use inside it. The insets indicate the location of the main maps in the Caribbean. The circle indicates the area enlarged in b. b) Leatherback nesting beaches at the West of the Park, Bahía de las Águilas and La Cueva.

Data collectionDuring the leatherback nesting season (from March to August, with a peak in April-June), we carried out night and daytime surveys at Bahía de las Águilas and La Cueva nesting beaches (See Revuelta et al. 2012 for an exhaustive description). When we found a clutch, we camouflaged the tracks of females and recorded GPS coordinates; we also recorded distance from nest to high tide line, the beach sector (Bahía de las Águilas and La Cueva), and the zone of the beach where the clutch was located (open sand, or within the dune vegetation).

We studied a total of 109 clutches laid on the beaches of Bahía de las Águilas and La Cueva over four nesting seasons (2007-2010). We excavated clutches by hand one to three days after hatchling emergence was detected. Clutch size was defined as the total number of eggs (hatched and unhatched yolked eggs)·clutch-1. Hatching success was calculated by dividing the number of eggs hatched by the clutch size expressed as percentage (Miller 1999). Leatherback turtle clutches have a high frequency of eggs without yolk commonly referred to as yolkless eggs which were also counted during the excavation but are not accounted for in clutch size.

Statistical analysisWe undertook detailed data exploration before any statistical analysis following Zuur et al. (2010). When the underlying question is to determine which covariates are driving a system, the most difficult aspect of the analysis is probably how to deal with correlation between covariates. We used variance inflation factor (VIF) and multi-panel scatterplots to test for covariate collinearity. We used boxplot and Cleveland plots for outlier detection, and two-way relationships were studied through multi-panel scatterplots between the response variable (percent hatching success) and each covariate.

Restricted maximum likelihood with linear mixed models was used to investigate the effect of environmental and temporal vari-ables on the hatching success. The response variable (percent hatching success) was logit transformed and analysed as a function of the explanatory variables considered likely to be important determinant predictors of


the hatching success for sea turtles in previ-ous studies (Wallace et al. 2004; Ditmer and Stapleton 2012; See Table 7.1 for descrip-tion). We included year as random factor in our model. It should be noted that we excluded 2010 data from statistical analy-ses because some of the predictor variables were not recorded in this year due to field-work limitations. To identify a suitable and parsimonious approximating model, we first developed a series of alternative mixed effects models that included different com-binations of the explanatory variables using stepwise process. Model selection was based on ΔAIC values lower than 2, calculated

as the difference between the AIC values for each model and the model with lowest AIC, and model weights (wi) (Burnham and Anderson 2002). Model weight is a relative index for model’s likelihood against any other model in the set and it was also used to calculate the relative importance of a variable by summing the weights of all the models that include that variable (Burnham and Anderson 2002).

All analyses were performed using the lnme package (Pinheiro et al. 2011) in R (version 2.14.0) for the linear mixed models.

Table 7.1. List of explanatory variables included in mixed models to model the leatherback hatching success in the JNP. The name of each variable entered into the models, description of each one, type and levels of the categorical variables are presented.

Name Variable description Type of variable Levels of variables

LOCATION

DIST Distance (m) to high tide line Continuous

ZONE Beach zone Categorical Open sand/Vegetation

SECT 2 sections separated by a rocky zone Categorical Bahía de las Águilas/ La Cueva

TEMPORAL

YEAR Breeding season year Categorical 2007, 2008, 2009

JULIAN Date of clutch laying Continuous

REPRODUCTIVE

ClutchSZ Clutch size (number of yolked eggs) Continuous

YLS Number of yolkless eggs Continuous

ID Incubation duration (number of days between egg laying and the time of first hatchling emergence)

Continuous


RESULTS In Bahía de las Águilas beach, mean ± SD hatching success of leatherback turtles clutches across years was 54.9 ± 15.5% (n = 6), 83.5 ± 20.9% (n = 19), 66.1 ±

We analyzed and modeled the hatching success of 64 clutches laid by leatherback turtles spanning three nesting seasons (2007, 2008 and 2009). No correlation was found between any pair of variables which were, therefore, included in the model testing the effect of 7 biotic and abiotic covariates (see Table 7.1). Of the 25 candidate models considered, five were retained in the models selected by AIC (Table 7.2). The hatching success of leatherback turtles appeared to be mainly driven by the effects of characteristics of beach sector, the incubation duration and date of lay, as evidenced by the retention of these variables in the most parsimonious model (AIC= 211.32) and in all five models

with ΔAIC values lower than 2 (Table 7.2). The relative importance of beach sector (Bahía de las Águilas and La Cueva), incubation duration and date of lay was 0.99 and the relative importance of clutch size, beach zone (open sand, or within the dune vegetation) and distance to high tide line, was 0.75, 0.35 and 0.24, respectively. Although the number of yolkless eggs was initially included in the models, this variable was not significant.According to the best model, hatching success of clutches incubated in the beach sector of La Cueva was higher than in those incubated in Bahía de las Águilas (Table 7.3, Figure 7.2b).

24.9% (n = 16) and 76.1 ± 20.3% (n = 40) in 2007-2010, respectively). In La Cueva beach, mean ± SD hatching success was 81.5 ± 16.3 (n = 6) in 2007, 79.1 ± 12.0 (n = 17) in 2009 and 84.2 ± 9.6 (n = 5) in 2010 (Figure 7.2a). No clutches were studied in 2008 in La Cueva beach.

Figure 7.2. a) Hatching success of leatherback turtles in Bahía de las Águilas (white) and La Cueva beach (grey). a) Actual data across years during the study period. No clutches were incubated in La Cueva in 2008. b) Fitted values by locality predicted by the best model (see Table 7.2) compared with actual values. Numbers above bars indicate sample size.


Table 7.2. List of models with better fit to the data are presented sorted according to Akaike information criterion (AIC) values. Five (out of 25) models with values of ΔAIC lower than 2 are presented. Weights of evidence in support of a particular model given the data (w) are also listed.

Models AIC ΔAIC w

HS ~ SECT + ID + JULIAN + ClutchSZ + YEAR (random factor) 211.32 0.00 0.28

HS ~ SECT + ID + JULIAN + ClutchSZ + DIST + YEAR (random factor) 211.60 0.28 0.24

HS ~ SECT + ID + JULIAN + ClutchSZ + ZONE + YEAR (random factor) 211.78 0.46 0.23

HS ~ SECT + ID + JULIAN +YEAR (random factor) 212.98 1.66 0.12

HS ~ SECT + ID + JULIAN + ZONE +YEAR (random factor) 213.12 1.79 0.12

Table 7.3. Estimates and standard errors (SE) from the selected model to explain the factors affecting hatching success of leatherback turtles in JNP. Model terms are described in Table 7.1.

Variable Estimate SE

Intercept 10.227*** 2.903

SECTOR La Cueva 0.711** 0.329

ID -0.088** 0.036

JULIAN -0.016** 0.006

ClutchSZ -0.017* 0.009

Significance codes: *** P = 0-0.001; ** P = 0.01-0.05; * P = 0.05-0.1


In addition to beach sector, hatching suc-cess was strongly influenced by incubation duration (Table 7.3); longer incubation du-rations resulted in lower hatching success (Figure 7.3a). We also found date of lay af-fecting hatching success, with clutches laid earlier in the nesting season having higher hatching success (Table 7.3, Figure 7.3b). The model also showed that increasing clutch size had a negative effect on hatching success (Table 7.3, Figure 7.3c).

Figure 7.3. Effect of a) incubation duration, b) julian day (date of lay) and c) clutch size (number of eggs incubated), on hatching success of clutches at western beaches of JNP. Shown are fitted values of the best model. Trends lines are cubic smoothing splines fitted by generalized cross-validation.

DISCUSSIONOur results show that clutches of leatherback turtles in the western beaches of the JNP presented unusually high hatching success (75.2%) for this species in the Caribbean (~ 50% see Eckert et al. 2012 for review). These beaches are of high value for the recovery of this threatened species in the Dominican Republic, taking into account the high level of egg take and the difficulties in protection in other beaches in the country (Revuelta et al. 2012). Studies on other beaches in the Caribbean region have suggested that bacterial and fungal infections (Patiño-Martínez et al. 2011), egg predation by ghost crabs, ants and other arthropoda (Maros et al. 2003; Caut et al. 2006) or non-natural predators such as feral dogs (Ordoñez et al. 2007) are important factors reducing hatching success. Additionally, the invasion of the clutch by vegetation roots, particularly from the species Ipomoea pes-caprae, may also have adverse effects on hatching success (Caut et al. 2010; Conrad et al. 2011). Although no quantitative data were recorded, fungal attacks (according to description in Chan and Solomon 1989) were detected in a very few clutches. The ghost crab (Ocypode quadrata) is present at the beaches; however, we did not observe egg predation by crabs, nor did we observe predation by ants or dogs or inundation by roots. Moreover, Ipomea pes-caprae is not present in the western beaches of the JNP. Taken together the absence of these factors likely contributes to enhanced hatch rate at site. Hence, our results highlight the need to preserve the relatively pristine environments with native beach vegetation to maintain the current levels of hatchling production.


Although both sectors studied presented similar nest density, sand and vegetation, clutches at La Cueva beach had even higher hatching success than at Bahía de las Águilas, possibly due to a greater slope in the former. Apart from beach sector, the study of factors affecting hatching success revealed that incubation duration and date of lay affected hatching success. The lower hatching success found in those clutches with longer incubation durations might be associated with process of washover which could cause temperature decrease inside the clutch, thus affecting embryo development (Houghton et al. 2007; Caut et al. 2010). Furthermore, the Caribbean hurricane season starts in June when many clutches are still incubating on the beach. Hence, it is possible that the low hatching success found in clutches laid later in the season might be associated with variations on incubation temperatures due to more violent wave regimes or flooding (Santidrián Tomillo et al. 2009). Future research should explore the roles of other factors that could not be recorded in this study such as nest depth, sand structure and incubation temperature, as well as individual-based reproductive variables that significantly affected hatching success in other nesting areas (Rafferty et al. 2011; Perrault et al. 2012). Since no spatial factors such as beach zone or distance to the shoreline seemed to affect hatching success, in situ clutch protection seems to be sufficient to preserve hatchling production in these beaches, avoiding the potentially negative effects of clutch relocation on hatching success (Pintus et al. 2009).

Current management strategies include surveillance of beaches and clutches by

the park rangers in Bahía de las Águilas. However, the Dominican government lacks the necessary human and economic resources to effectively manage protected areas and offenses to environmental laws are frequent in these beaches, as happens in other protected areas of the country (García and Roersch 1996; Kerchner et al. 2010). Given the exceptional value of hatching success in these beaches and current and potential threats affecting leatherback nesting beaches in the country, additional effort in regulation and management of this protected area is needed. Potential actions that can be taken to improve management strategies should include increased beach surveillance by rangers during the leatherback nesting season, banning access of motorized vehicles to the beaches as well as the use of lights and campfires. Government policies should include increasing awareness to facilitate a reduction in pollution, preserve the environment, and protect the endangered species of the Park (Choi and Eckert 2009).

An added threat to the lack of enforcement is the strong pressure to bring mass tourism to this area (both in the protected and non-protected beaches) in the form of high capacity resorts. Our study is particularly relevant in relation to La Cueva beach; this sector harbors 20% of the total clutches laid at western beaches of the Park and demonstrated the highest hatching success. However, it is less protected because it is located outside the Park limits, in the buffer zone. This means a higher level of threat as has been observed elsewhere in the country (Kerchner et al. 2010). Allowing coastal development in


the buffer zone would increase detrimental effects of human activities that would impact no only this zone but also the areas inside the Park. We strongly recommend including this beach inside the limits of the park thus conferring the same level of legal protection and surveillance as Bahía de las Águilas.

CONCLUSIONSAlthough Bahía de las Águilas is one of the best preserved areas in the Dominican Republic, hosting endemic species, it is threatened with increasing interest in development with plans including building of resorts and tourist facilities. Our study highlights the importance of this site and the neighbouring beach of La Cueva for leatherback turtle reproduction and the need of establishing management measures for a successful in situ conservation of this threatened species in the country. These protection measures would be also beneficial for the preservation of other species.

ACkNOWLEDGEMENTSWe are very grateful to fieldwork assistants, especially to the JNP rangers for their col-lective conservation effort and for assis-tance with data collection for this project. We extend our thanks to the Dominican and international students, volunteers and local villagers who helped in different tasks of the project. The present study is includ-ed as part of a conservation project fund-ed by: the Spanish International Cooper-ation Agency (AECI, projects: A/2991/05 and A/5641/06), the Spanish Ministry of Education and Sciences (CGL2006-

02936-BOS), the General Foundation of the University of Valencia, and the Euro-pean Union (Marie Curie grants, FP6 & 7). We specially thank the support of the Dominican NGO Grupo Jaragua and of The Nature Conservancy (TNC). The project counted on the support of the Do-minican Environmental authorities (office of Protected Areas and Biodiversity of the Environmental Ministry). J. Tomás and J.A. Raga are supported by project Pro-meteo/2011/40 of Conselleria de Edu-cacio (Generalitat Valenciana) and project CGL2011-30413 of the Spanish Ministry of Sciences and Innovation.


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8. CHAPTER VIII: EVALUATING THE IMPORTANCE OF MARINE PROTECTED AREAS FOR THE CONSERVATION OF HAWkSBILL TURTLES (ERETMOCHELYS IMBRICATA) NESTING IN THE DOMINICAN REPUBLIC

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ChApTEr VIII: EVALUATING THE IMPORTANCE OF MARINE PROTECTED AREAS FOR THE CONSERVATION OF HAWkSBILL TURTLES (ERETMOCHELYS IMBRICATA) NESTING IN THE DOMINICAN REPUBLIC

CHAPTER VIII:

EVALUATING THE IMPORTANCE OF MARINE PROTECTED AREAS FOR THE CONSERVATION OF HAWkSBILL TURTLES (ERETMOCHELYS IMBRICATA) NESTING IN THE DOMINICAN REPUBLIC

Ohiana Revuelta1, Lucy Hawkes2, Yolanda M. León3,4, Brendan J. Godley2, Juan A. Raga1 and Jesús Tomás1

__

Endangered Species Research (Submitted)

__1 Cavanilles Institute of Biodiversity and Evolutionary Biology, University of Valencia, P.O. Box

22085, E-46071 Valencia, Spain.2 Centre for Ecology and Conservation, School of Biosciences, University of Exeter Cornwall Campus,

Penryn, UK.3 Instituto Tecnológico de Santo Domingo, Urb. Galá, Santo Domingo, Dominican Republic.4 Grupo Jaragua, El Vergel, Santo Domingo, Dominican Republic.

08

Chapter VIII: Evaluating the importance of Marine Protected Areas for the conservation of hawksbill turtles (Eretmochelys imbricata) nesting in the Dominican Republic I 173

ABSTRACTUnderstanding spatial and temporal habitat-use patterns to protect both foraging and breeding grounds of species of concern is crucial for successful conservation. Saona Is-land in Del Este National Park (DENP), southeastern Dominican Republic (DR), hosts the only major hawksbill (Eretmochelys imbricata) nesting area in the DR (100 nests.year-1 ± 8.4 s.d., range: 93-111), with the population having been critically reduced through hunting. We satellite tracked nine hawksbill turtles and present analyses of their core-use areas with respect to Marine Protected Areas both in their internesting and foraging ar-eas. Kernel utilization distributions indicated that during the internesting period all turtles remained close to their nesting beaches in small home ranges in the territorial waters of DR, mostly over the continental shelf (<200 m depth). Common core-use areas were located inside the DENP and 82.7 % of all locations were within the DENP. At foraging areas, only 23 % of locations were inside MPAs either in waters of the DR and in waters of Bahamas, Nicaragua and Honduras. Our results highlight that the protected areas of the DR are key for hawksbill conservation, and enforcement of existing legislation of the protected areas in the country is key. The present study also corroborates that the waters off Nicaragua and Honduras are exceptionally important foraging areas for hawksbills in the Caribbean, showing the turtle’s vulnerability in these waters.


INTRODUCTIONMany threatened marine vertebrate spe-cies are of conservation concern as the re-sult of a range of past and ongoing threats (Read et al. 2006; Hoffman et al. 2011); and the establishment of marine protect-ed areas (MPAs) has been promoted as a key management measure for their con-servation (Halpern 2014). Management and conservation effectiveness of MPAs requires full knowledge about life history stages of the species they are intended to conserve (Edgar 2011). For highly migra-tory marine species this means knowledge about their migration routes as well as their movements and use of foraging and breed-ing grounds (Blanco et al. 2012; Costa et al. 2012). Owing to a revolution in the use of location technologies for tracking the at-sea movements of a range of marine spe-cies (‘biologging’, Bograd et al. 2010; Hart and Hyrenbach 2010; Hammerschlag et al. 2011), we now know a great deal about the movements and behaviour of many for-merly cryptic marine species and arguably the most about the marine turtles (Godley et al. 2008; Maxwell et al. 2011; Scott et al. 2012; Schofield et al. 2013).

The hawksbill turtle (Eretmochelys imbricata) is the most tropical of all marine turtles, dis-tributed in tropical and subtropical areas of the Atlantic, Pacific and Indian oceans, and currently is listed globally as critically en-dangered (IUCN 2013). In the Caribbean, this species is a major predator in coral reef ecosystems, with sponges constituting the main component of its diet (Meylan 1988; León and Diez 1999). Numerous studies have suggested the key role of this species

in maintaining the structure, ecology and evolution of coral reefs (León and Bjorndal 2002; Bjorndal and Jackson 2003). During the nesting season, which usually spans sev-eral months, females are thought to return repeatedly to the same beach to lay a vari-able number of clutches, thus presumably remaining close to the nesting beach (Mar-covaldi et al. 2012; Walcott et al. 2012). After breeding, hawksbill turtles are then thought to migrate away from the nesting beach to foraging grounds (Meylan et al. 2011). Migratory movements of this species in the Caribbean, tracked using satellite te-lemetry, show a broad range of movements, including turtles remaining in waters proxi-mate to the nesting beaches and others mi-grating to foreign waters many hundreds to thousands of kilometres away (van Dam et al. 2008; Horrocks et al. 2011; Hawkes et al. 2012; Moncada et al. 2012).

The Dominican Republic (DR) in the East-ern Caribbean, hosts regionally significant numbers of nesting hawksbill turtles (Re-vuelta et al. 2012) but current nesting is largely restricted to protected areas. Saona Island in Del Este National Park (DENP), southeastern DR, hosts the last major nest-ing area in the country (~100 nests per year; Revuelta et al. 2012). Although nest-ing beaches are in a protected area, there has been a marked reduction in the abun-dance of this species and egg take is still a major threat at these beaches (Revuelta et al. 2012; 2013).

Here we reanalyse these tracking data with respect to (1) behaviour of hawksbill fe-males during the internesting period and, (2) an assessment the level of protection af-forded to these turtles in their internesting and foraging areas.



Tagging area Surveys for nesting hawksbill turtles were carried out in the Jaragua National Park (JNP, southwest DR, 17816′N-71533′W; Figure 8.1) and, Del Este National Park (DENP, southeastern DR, 18807′N-68844′W; Fig-ure 8.1). Both parks were added to the UN-ESCO World Heritage list in 2001. JNP covers 1,374 km2, of which 905 km2 com-prises coastal areas and hawksbill turtles nest at the westernmost beaches, Bahía de las Águilas (4.4 km long) and La Cueva (2.5 km). The park receives c.a. 20,000 visitors per year (Wielgus et al. 2010). DENP is the second largest coastal national park in the DR, and comprises some of the southeast-ern DR and Saona Island at approximately 110 km2. Hawksbill nesting activity is con-centrated on 12 sandy beaches (altogether comprising a total of 15 km) interspersed with rocky areas (Revuelta et al. 2013). The adjacent neritic zone is characterized by a wide continental shelf comprising fringing reefs and rocky bottom communities mixed with seagrass beds. Despite DENP’s desig-nation as a national park, the government of the DR allows small-scale artisanal fish-eries to occur within its boundaries, mainly for local consumption. However, overfish-ing has affected the area with populations of invertebrates (conch and spiny lobster) and reef fish highly depleted (Chiappone et al. 2000). In addition, Saona Island is one of the most popular tourist destinations in the DR with up to one thousand visitors a day (MacLeod 2001) resulting in a heavy traffic of motor boats, especially at the west part of the island.

Figure 8.1. Study area. a) Map of DR indicating hawksbill nesting and foraging areas at the protected areas of JNP (southwest DR) and DENP (southeast DR, including Saona Island) where turtles were tagged. B) Hawksbill foraging areas at the Western Caribbean. Boundary lines for MPAs in the Caribbean Sea are shown. HA: Haiti; HO: Honduras; NC: Nicaragua; CRC: Costa Rica.

Marine turtle tracking and satellite data filtering We satellite-tagged a total of nine nesting hawksbill turtles, eight on Saona Island in August-September 2008 and Septem-ber-December 2009; and one (Ei8) in JNP in September 2009. Turtles ranged in size from 81.0 to 94.0 cm CCL (Mean ± SD: 87.3 ± 4.4 cm; Table 8.1). No turtles had previous tags. Tags were Wildlife Comput-ers SPOT5 tags (n = 5) and Sirtrack Ki-wisat 101 tags (n = 4). To attach the units, we detained each turtle inside a portable


wooden corral following nesting or nesting attempts, intercepting them in their way back to the sea. The carapace of each tur-tle was prepared by scrubbing to remove epibionts, sand, and cleaning with acetone. The units were attached to the second ver-tebral scute of the turtles’ carapace with 2-part epoxy and covered with a layer of anti-fouling paint (Blumenthal et al. 2006). Before attaching the transmitter, each turtle was measured (curved carapace length) and tagged with small metal Inconel tags to the front flippers. Once tagging was complete we removed the corral, allowing the turtle to return to the sea.

For all transmitters, data were downloaded from the ARGOS satellite system and sub-sequently analyzed with the satellite track-ing and analysis tool (STAT, Coyne and Godley 2005) to archive and filter location data. For each reported location, ARGOS calculates a measure of accuracy using six “location classes” (LC): 3, 2, 1, 0, A, and B). As in previous hawksbill telemetry studies (Cuevas et al. 2008; Gaos et al. 2012) the majority of our LCs were categorized as B, thus we considered the positions LC: 3, 2, 1, A and B to avoid loss of relevant location data. These LCs were retained and filtered to remove biologically unrealistic speeds (>5 km h-1; see Luschi et al. 1998), turning angles (<25° were removed) and elevations (>0 metres above sea level).

Turtle’s core-use areas and distribution within MPAs To minimize autocorrelation in spatial analysis we generated mean daily locations for each turtle from the filtered locations (Hart et al. 2010, 2013). To determine core-use areas, foraging and nesting data were extracted using displacement plots and assessed separately (Hawkes et al. 2012). Home range size was estimated using min-imum convex polygons (MCP), a non-sta-tistical measure which encapsulates the area used by an individual within a polygon formed by joining the outer-most sighting positions (Burt 1943). MCP is a simple cal-culation that allows for comparisons be-tween studies (Hooge et al. 1999), but is unable to describe fine-scale movements and preferred area used within the poly-gon. It may also be inflated by inaccurate yet plausible data outside the true area of utilisation (Laver and Kelly 2008). There-fore, core-use areas were identified using fixed kernel density estimation (KDE) with individual kernel contours delineated using a smoothing factor (h) of 1.5 for internest-ing and 2.5 for foraging. Density distribu-tions were represented on the maps by the 50% and 90% utilization distribution (UD) contours. We used a 90% KDE to estimate the overall home foraging and internesting range of a turtle and a 50% KDE to repre-sent the core area of activity (Powell 2000). For turtles tracked through two subsequent nesting seasons KDE was calculated sepa-rately for each internesting period. Follow-ing the methodology of Hart et al. (2013) we did not estimate KDEs for turtles that transmitted for less than 20 days (i.e., had less than 20 mean daily locations).


Common use areasCommon core-use areas were generated for all turtles combined (i.e. where multiple tur-tles spent time during the internesting and foraging periods). These areas were deter-mined using individual 90% and 50% ker-nel-density estimates (KDEs).

To analyse the location of turtles with respect to MPAs, filtered turtle location data were overlaid on the World Database on Protect-ed Areas (www.wdpa.org). Site fidelity was quantified using a residency index (Mason and Lowe 2010) calculated by dividing the number of days a female was detected with-in MPAs’ boundaries by the number of days the female was monitored (i.e. internesting period). Values range from 0, indicating no residency, to 1 indicating a high degree of residency. All spatial analyses were carried out in ArcGIS 10 (ESRI 2010).

Characterization of turtle’s movements during internesting period In the present study, and similar to meth-odology used by Rees et al. (2010), Tucker (2010) and Maxwell et al. (2011), nesting ac-tivity was extracted from the tracking data by evaluating the following criteria, classing a nesting event as having occurred when: (a) locations were within 1km of the coast-line; (b) the turtle made directed on shore movement; (c) movements occurred within the known internesting intervals for hawks-bill turtles (~15 days, Beggs et al. 2007); (d) high-Argos LCs 3, 2, and 1 within a short time span.

RESULTSOverall transmission success rate (mean number of locations received per day, re-ferred to as ‘MDL’ from here on) ranged between 1.8 and 5.6 locations per day (SD range 1.3 to 2.7 locations per day) during the internesting period, and 1.0 to 4.2 (SD range 1.2 to 2.0 locations per day in foraging grounds; Table 8.2). Transmitter duration was variable: Turtle Ei1 was only tracked during the internesting period; turtles Ei4 and Ei8 departed immediately following device attachment and were therefore only tracked during the foraging period; the six other turtles (Ei2, Ei3, Ei5, Ei6, Ei7 and Ei9) were tracked during both internesting and foraging (Table 8.1). Transmitters attached to three turtles (Ei2, Ei3 and Ei9) functioned for particularly long periods, permitting in-sight into multi-year space use. These three turtles were tracked from arrival at their for-aging grounds until their departure to breed and nest again in the DR.

Insight into internesting period behaviourFor the internesting period, we analysed a total of 370 tracking days from the seven turtles for which we had internesting data (see above). Individual tracking durations during the internesting period ranged from 0 to 64 days (Mean ± SD: 37.4 ± 22.7 days, Table 8.1). Excluding turtles Ei4 and Ei8 as well as turtle Ei9 in its first nesting season that immediately left coastal waters after laid the last clutch, we estimated that hawksbill females migrate from the internesting area after laying their last clutch of the season a mean (±SD) of 4.3 ± 5.8 days later (range: 0-14 days; Table 8.1). Twenty nesting events were inferred from seven turtles (including


two turtles that were recorded nesting in two different years) (Table 8.1). We estimated a mean (±SD) time between nesting events (in-ternesting interval) of 15.5 ± 0.9 days (range: 14-17 days). Turtles E1, Ei2 and Ei3 did not successfully nest at the time of transmitter deployment, thus internesting interval was determined from the first inferred nesting date instead. Mean (±SD) minimum clutch frequency of 7 females for which we have records was 2.9 ± 0.6 clutches (range: 2 to 4 clutches; Table 8.1).

During the internesting period all turtles remained in the territorial waters of Do-minican Republic, mostly over the conti-nental shelf (<200 m depth; Figure 8.2). Turtles spent most of their time in areas

characterized by relatively shallow waters (KDE 90% for all turtles over water less than 100 m deep). Turtles Ei2 and Ei5 made excursions beyond the shelf (e.g. to the 1000 m isobath, Figure 8.2b and 8.2d). Turtles were usually located 1.4 to 4.3 km from the coast (SD range: 2.1-5.7 km) and mean maximum distance from the coast was 22.4 km (SD 9.2; range: 13 to 39 km; Table 8.1). Analyses of the turtles daily lo-cations show that during the internesting period, of the 370 total tracking days, 306 (82.7%) were within the DENP’s borders. Excluding turtle Ei5, turtle’s residency in-dex ranged from 0.65 to 1.0 (mean ± SD: 0.84 ± 0.1) indicating a high degree of use of the protected area.

Figure 8.2. Eretmochelys imbricata. Individual internesting areas in DENP indicated by minimum convex polygons (MCP) and 90% (black) and 50% (light grey) utilization distributions (UDs) for (a) turtle Ei1; (b) turtle Ei2; (c) turtle Ei3; (d) turtle Ei5; (e) turtle Ei6; (f) turtle Ei7 (g) turtle Ei9; (h) Common core-use area of seven turtles tracked during the internesting period identified using 50% and 90% utilization distributions (UDs). For turtles Ei2 and Ei3 two MCP areas: 1st internesting period (black) and 2nd internesting period (light grey) are shown. Ei2 and Ei3 UDs depict overlap zones 50% and 90% KDEs of two internesting periods. DR: Dominican Republic. SI: Saona Island. Coral reef ecosystems are indicated by hatched areas and marine protected areas by dashed black lines. Stars indicate tagging location. Dashed light grey lines: 200 m and 1000 m bathymetric contour.


Table 8.1. Summary information of Saona nesting hawksbill behaviour and habitat utilisation during internesting period. CCL: curved carapace length. IN: internesting. IN tracking duration: defined as time from tagging date until the last internesting location. Inferred nesting dates: nesting emergences inferred from satellite-tracking data during the IN interval. IN interval: number of days between two consecutive nesting events. Turtles Ei3 and Ei9 stopped sending signals during their second internesting period in 2011 thus we could not estimate time in the IN after the last nest. Turtles Ei4, Ei8 immediately left coastal waters; therefore no internesting interval and clutch frequency could be estimated for these turtles.* Turtles tracked during two successive nesting seasons.

Turtle Tag date CCL(cm)

IN trackingduration

(days)

Inferred nesting

dates

IN interval(days) Mean

± SD

Clutch frequency

Days in INafter last

nest

Swim speed(km h-1) Mean ± SD

Max distanceoffshore

(km)

Ei1 11/08/2008 84 34 15/08/200830/08/200814/09/2008

15.0 ± 0.0 3 1 1.4 ± 1.2 17

Ei2* 28/08/2008 90 60 16/09/200829/09/200814/10/2008

14.0 ± 1.4 3 14 1.5 ± 1.3 28

Ei2 — — — — — — — — 33

Ei3* 29/08/2008 90 48 15/09/2008 30/09/2008 14/10/2008

15.3 ± 1.4 3 1 1.2 ± 1.3 25

Ei3 — 52 24/07/201110/08/2011

17.0 2 — 0.7 ± 0.8 13

Ei4 19/09/2008 94 0 — — — 0 — —

Ei5 30/09/2008 92 28 15/10/2008 15.0 2 13 2.4 ±2.7 39

Ei6 27/10/2008 81 51 28/11/200814/12/2008

16.0 3 2 1.4 ±1.6 17

Ei7 01/11/2008 84 37 18/11/200803/12/2008

16.0 ± 1.4 3 4 1.4 ± 1.5 16

Ei8 07/08/2009 84 0 — — — 0 — —

Ei9 01/09/2009 87 64 02/07/2011 18/07/2011

02/08/2011 18/08/2011

15.7 ± 0.6 4 — 0.8 ± 1.0 14


during internesting there was no significant variation in home range size among the 7 hawksbill turtles tracked during their in-ternesting period (Figure 8.3), with large portions of MCPs and KDE overlapping (Figure 8.2). Turtles Ei2 and Ei3 showed internesting habitats with similar size and location during their two nesting tracked seasons (Figure 8.2b and 8.2c). There were no correlations between turtle size and the number of days the turtles spent in in-ternesting areas (Pearson’s t = -0.7, df = 5, p= 0.5) and between turtle size and the in-ternesting area size (Pearson’s t = -1.2, df = 5, p= 0.3).

Core use internesting areas We calculated MCPs for these 7 turtles, two of which transmitted data for a second nesting season, resulting in 9 MCPs (tur-tles Ei2 and Ei3; Table 8.2 and Figure 8.2). Internesting areas occupied by the turtles ranged from 51.5 to 644.3 km2 (mean ± SD: 254.5 ± 173.5, Table 8.2). The 90% KDEs (n = 8) for the 6 turtles for which in-ternesting core areas were calculated were much smaller and ranged from 19.6 to 86.0 km2 (mean ± SD: 49.1 ± 20.1 km2, Table 8.2); and the mean ± SD 50% KDEs area was 13.7 ± 4.9 km2 (5.4 to 21.1 km2, Table 8.2). With the exception of one turtle (Ei2b)

Figure 8.3. Home range size among the 7 hawksbill turtles tracked during their internesting period in the Dominican Republic. MCP: Minimum Convex Polygons; KDE: Kernel Density Estimation (see Material and Methods section for descriptions).


Table 8.2. Transmission success (mean number of locations received per day, mdl), Minimum convex polygon (MCP) and Kernel density estimation (KDE) for hawksbill turtles in their internesting and foraging grounds. BH: Bahamas; CO: Colombia; DR: Dominican Republic; HO: Honduras; NC: Nicaragua. Turtle Ei1 stopped transmitting before arrive the foraging ground. Turtle Ei2 was tracked during two nesting and foraging seasons. Ei3 was tracked during two nesting seasons and one foraging season. Turtles Ei4 and Ei8 departed immediately following device attachment. Turtle Ei5 had less than 20 mean daily locations (n= 16) during the internesting period for KDE analysis. Minimum Convex Polygon sizes in foraging grounds also reported in Hawkes et al. 2012.

Internesting period Foraging period

Turtle mdl MCP(km2)

90% kDE(km2)

50% kDE(km2) mdl MCP

(km2)90% kDE

(km2)50% kDE

(km2)Maritime

boundaries

Ei1 3.1±2.0 266.6 32.5 12.2 — — — — —

Ei2a 1.9±1.8 321.1 57.9 15.1 2.2±1.4 5007 173.1 21.4 NC/HO

Ei2b — 644.3 55.6 14.2 — 38 14.2 3.9 NC

Ei3a 3.8±2.6 325.4 39.9 11.8 2.5±1.5 25561 257.1 24.9 NC/COL

Ei3b — 51.5 19.6 5.4 — — — — —

Ei4 — — — — 1.8±1.2 9330 201.2 45.3 NC/HO

Ei5 2.1±1.8 109.5 — — 1.0±1.8 314 24.9 13.2 DR

Ei6 2.3±1.3 184.7 57.9 19.2 2.0±1.4 1704 63.6 6.9 DR

Ei7 1.8±1.3 243.0 86.0 21.1 1.3±1.2 3618.8 57.2 21.1 BH

Ei8 — — — — 2.4±1.4 252.1 46.8 4.1 HO

Ei9 5.6±2.7 144.1 43.8 10.4 4.2±2.0 15522 297.8 34.5 NC/HO

Core use foraging areas and MPAsWe also calculated MCPs and 90% and 50% KDEs for turtles in their foraging grounds (Table 8.2). Foraging areas of five females (Ei2, Ei3, Ei4, Ei8, Ei9) were located along the waters off Honduras and Nicaragua (Figure 8.3a). In those waters there are two main protected areas, the Miskito Cays and the Seaflower MPA (part of the Seaflower Biosphere Reserve, belonging to Colom-bia). The Miskito Cays (27 km²) is an archi-pelago located off shore in the North-East-ern Caribbean coast of Nicaragua (Figure 8.1). The seagrass beds and reefs in the cays are among the Atlantics’s greatest forag-ing grounds for green (Chelonia mydas) and hawksbill marine turtles (Bjorndal and Bol-ten 2003). The Seaflower MPA, located in the south-western Caribbean sea (Figure 8.1) is the largest MPA in the wider Carib-

bean (65,000 km2) protecting mangroves, seagrass beds and the largest and most pro-ductive coral reefs in the region (Taylor et al. 2013).With the exception of turtles Ei3 and Ei4 with residency index of 0.02 and 0.91 respectively (Seaflower Biosphere Re-serve and Miskitos Cays respectively), the rest of the turtles were not located in pro-tected waters for any of their tracked forag-ing period (Figure 8.4a to 8.4e). After nest-ing in Saona, turtles Ei5 and Ei6 remained in waters of the DR within coastal reef ecosystems (Figure 8.5a and 8.5b). Turtle Ei5 stayed inside the JNP (residency index: 0.91; Figure 8.5a). Turtle Ei6 foraging area was located in waters adjacent to Bahía de las Calderas (southern DR coast) outside of marine protected areas (Figure 8.5b). Turtle Ei7 travelled to the northwestward to the Bahamas where its core use area was not in protected waters (Figure 8.5c).


Figure 8.4. a) MCPs of hawksbill turtles at Nicaragua and Honduras foraging area. (+): MCP of Ei2b (2012 foraging period). Core-use areas defined by 90% (black) and 50% (light grey) utilization distributions (UDs) plotted with adjacent MPAs of Miskito Cays and Seaflower (dashed black lines) for b) turtle Ei9 which occupied two separated areas during the same foraging period; c) turtle Ei8 d) turtle Ei4 and e) turtles Ei2a (2008 foraging period) and Ei3 (inset: UDs for turtle Ei2b during 2012 foraging period, which overlapped with UDs in 2008). Note different scales. Dashed light grey line: 200 m bathymetric contour.

Figure 8.5. Foraging core use areas defined by MCPs and 90% (black) and 50% (light grey) utilization distributions (UDs) of a) turtle Ei6 near shore of Las Calderas Natural Monument (south DR) b) turtle Ei5 in waters inside of JNP boundaries (Southwest DR); and c) turtle Ei7 in waters of Bahamas. Note different scales.


Common-use areasThe internesting common-use area 90% and 50% KDE was 81.7 km2 and 32.2 km2 respectively (Figure 8.2h). This com-mon-use area was situated on coral reefs at the eastern-most tip of the island and the individual home ranges for all seven turtles overlapped, showing commonality across turtles and years. Overall, the com-mon core-use area during nesting was situ-ated inside the DENP’s boundaries (Figure 8.2h). There was, however, no overlap in foraging grounds used by females tracked in the same year; but, there was overlap in resident areas utilised by one turtle that was tracked in two different foraging period (Ei2) used same location in both years (Fig-ure 8.4e).

DISCUSSION

Core use areas and MPAsHome range estimations provide knowl-edge of marine turtles’ core areas of ac-tivity, underscoring hotspots for their pro-tection (Scott et al. 2012; Schofield et al. 2013). The present study has permitted insight into the effectiveness of some exist-ing MPAs in the protection of the nesting hawksbills. It should be noted that the num-ber of tracked turtles studied in the present paper represent ca. 40% of the total annu-al nesting stock in the Dominican Repub-lic (Revuelta et al. 2012). Overall, during internesting intervals turtles were located within MPAs while during the foraging pe-riod they were mostly (78.0 % of total for-aging days tracked) outside of any MPA.

Our results reveal that during their in-ternesting period hawksbill turtles in Saona Island are mostly within the maritime lim-its of the DENP, highlighting the impor-tance of effective management of this area for their conservation. Although the spatial extent of the DENP offers a good opportu-nity to enhance the protection for this crit-ical breeding ground, turtles face multiple anthropogenic threats. Firstly, the expansion of tourism industry has increased the boat traffic around Saona Island, particularly in the west part, and also increasing the pres-sure on coral reefs due to direct pollution (Wilkinson 2000). Secondly, artisanal fisher-ies have essentially uncontrolled access to the park, resulting in the depletion of large reef fish, conch and lobster populations (Chiap-pone et al. 2000). In addition, illegal capture of adult turtles by illegal fisherman has also been documented (Revuelta et al. 2013).

Hawksbill turtles tagged in Saona show a range of migratory strategies, with some turtles remaining near nesting sites in the DR waters and others migrating to inter-national foraging grounds (Hawkes et al. 2012). Most of the turtles that migrated internationally foraged in waters off Nic-aragua and Honduras (n = 5), which has already been confirmed as a preferred foraging ground for hawksbills nesting in other Caribbean areas such as Costa Rica, Eastern Caribbean and Cuba (Troëng et al. 2005; Horrocks et al. 2011; Moncada et al. 2012). Of these 5 turtles just one (Ei4) spent a large proportion of the time inside a pro-tected area (Miskito Cays). However, the effectiveness of the protection in this area is questionable as a result of legal and ille-gal marine turtle fisheries (Bräutigam and Eckert 2006). The lack of protection in the


waters off Nicaragua and Honduras thus poses a potentially significant conservation problem for many Caribbean hawksbill nesting populations. In the Bahamas, there are eight marine national parks, and direct harvest of marine turtles is probably con-siderably less common. However, protected waters make up a minority of the 630,000 km2 exclusive economic zone of the Baha-mas (less than 1%) and thus it is not sur-prising that the turtle foraging here did so outside of protected areas.

Turtle Ei5 stayed in DR waters, spend-ing most of the time (91 % of the total days tracked) within JNP maritime limits, which supports previous studies about the importance of this area for marine turtle conservation in the country, not only as a foraging ground for juveniles and adults but also as a nesting area (León and Diez 1999; Revuelta et al. 2012). The second potential foraging ground in DR was lo-cated in the adjacent waters of the Nat-ural Monument Bahía de Calderas in the south coast of the country. This protected area encompasses around 15 km of sand dunes but does not yet incorporate the sea. Sporadic nesting by Chelonia mydas has also been detected in these beaches (Y.M. León pers. obs.). Despite the evident importance of this area for biodiversity conservation, and its protected status, it is threatened by the extraction of sand for commercial pur-poses and illegal construction by the hotel industry, as well as indiscriminate fishing activities (Perdomo et al. 2010). Such lack of enforcement of conservation actions in protected areas in the country prevents an effective protection of these critical habi-tats for marine turtles and is an urgent tar-get for improved conservation.

Internesting behaviourNesting hawksbills of Saona Island re-mained in the adjacent waters to their nest-ing beaches using small home range areas during internesting intervals. Core activity areas occurred in shallow waters main-ly within 200 m isobaths and associated with coral reefs. Our results support previ-ously described hawksbill internesting be-haviour observed in other internesting ar-eas throughout the Caribbean (van Dam et al. 2008; Marcovaldi et al. 2012; Walcott et al. 2012). Regardless of the nesting location on the beach, core-use areas were situated on coral reefs at the eastern-most tip of the island. Turtle preference for particular internesting areas has been related to ade-quate resources and quality of the habitat occupied (Hart et al. 2010). The abundance of fringing reef systems inside DENP’s ma-rine area likely accounts for hawksbill tur-tle affinity for this area. The abundance of sheltered resting sites could be an import-ant resource, helping nesting females to preserve energy during their reproductive period (Houghton et al. 2008; Hart et al. 2010). Although it may not the case every-where (e.g., Zbinden et al. 2007), we hy-pothesize that the huge number of tourist boats in the vicinity of nesting beaches in the west part of Saona, could lead to avoid-ance behaviour.

Internesting activity patterns for Saona hawksbill turtles were similar to results for hawksbill turtles satellite tracked during nesting periods in other locations. How-ever it should be noted that since tagging date was the only confirmed nesting event, nesting dates were inferred by turtle tracks (see Materials and Methods section), thus nesting dates and internesting intervals are


approximated. Internesting interval ranged between 14 to 17 days described for the species in closer areas (Beggs et al. 2007; van Dam et al. 2008). Tracking-derived minimum clutch frequency, which is unable to account for nests laid prior to transmit-ter attachment, ranged from 2 to 4 clutch-es, which is in accordance with conspecifics elsewhere (Witzell 1983). Previous research-ers have hypothesized that hawksbill turtles may not forage during internesting inter-vals and at the end of the nesting season, fe-males would need to migrate back to forag-ing grounds to restore their energy reserves (van Dam et al. 2008). The present study seems support this hypothesis since turtles remained in the internesting area only for a few days following their final nesting event. It is possible that the longer post-nesting pe-riods inferred for turtles Ei2 and Ei5 may have been due to repeat nesting that we were unable to identify, or because some foraging may take place.

Conclusions and conservation recommendationsIn this study, we describe DR hawksbill nesting turtles use of MPAs at nesting and feeding grounds, adding information to the use of MPAs by this species in the Caribbe-an region. This information highlights the significance of protected areas in the DR for internesting and foraging hawksbills, showing the need to enforce existing legis-lation of the protected areas in the country. The turtles use waters inside the protected areas of DENP and JNP year-round, ar-eas that are severely threatened by human activities. Hence, efforts must be increased to mitigate illegal fishing in the waters of these parks. In the case of waters around

Saona we recommend the creation of near shore zone of maximum protection that would enhance the protection of the rook-ery in this highly used area (i.e., by restrict-ing the boat traffic in this zone). Likewise, we propose the expansion of Las Calderas Natural Monument boundaries offshore due to its importance not only for turtles but also for other species and ecosystems, as well as enhanced enforcement of existing regulations in the Monument. The present study also corroborates that the waters off Nicaragua and Honduras are exceptionally important foraging areas for hawksbills in the Caribbean. We recommend that MPAs for marine turtle conservation in the region should be reassessed as a priority.

ACkNOWLEDGEMENTSFunding for the transmitters came from the JM Kaplan Foundation award to the World Wildlife Fund, Canada, and from different European institutions: the Spanish International Cooperation Agency (AECI, projects: A/2991/05 and A/5641/06), the Spanish Ministry of Education and Sciences (CGL2006-02936-BOS and CGL2011- 30413) and the General Foundation of the University of Valencia. The project received funding also from the European Union (Marie Cu-rie grants, FP6 and FP7). L.A.H. was sup-ported by the WWF MacArthur project ‘Developing an approach to adaptation to climate change in the marine turtles: the hawksbill turtle as an indicator species’ to C. Drews. Work was carried out under permit in the DR. L.A.H. is supported by a Biotechnology and Biological Sciences Research Council (BBSRC) post-doctoral


fellowship. J.T. and J.A.R. are supported by project Prometeo/2011/40 of Con-selleria de Educacio (Generalitat Valenci-ana) and project CGL2011-30413 of the Spanish Ministry of Economy and com-petitiveness. B.J.G. is supported by the Darwin Initiative, Natural Environment Research Council (NERC) and the Penin-sula Research Institute for Marine Renew-able Energy (PRIMaRE).


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CONCLUSIONS©

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Conclusions I 193

CONCLUSIONSThe present PhD Thesis analyses the conservation status of marine turtle nesting rookeries in the Dominican Republic. The following section summarizes the main findings and conclusions.

1) This study represents the first detailed assessment of the status of marine turtles nesting in the Dominican Republic, based on surveys during the period 2006-2010, and compares the results with previous studies. Leatherback, hawksbill and green turtles are still nesting in the country; how-ever, unlike previous studies, no loggerhead turtles nesting event was detected during this period.

2) Currently, marine turtle nesting activity is concentrated in the protected areas of Ja-ragua National Park and Saona Island (Del Este National Park). Nesting events out of these two areas are sporadic. The results of this study suggest that there has been a pro-found reduction in the abundance of the marine turtle species in the country since the 1980s.

3) The Jaragua National Park, in the south-west of the country, hosts the highest num-ber of clutches per year of leatherback turtles (mean 126.4 ± SD 74.1, range: 17-210), with a total of 632 clutches recorded during the period studied. Nesting season of leatherback turtles extends from March to August, with most emergences occurring in April-June. The estimated annual num-ber of leatherback nesting females varied from 3 to 33 turtles. The size (mean CCL = 147.4 cm) of leatherback turtles in the Do-minican Republic is smaller than the glob-

al mean size, which could be the sign of a recovering population with a high propor-tion of neophyte females. Sporadic nesting of hawksbills and green turtles was also re-corded in this area.

4) Saona Island, in the southeast of the country, is the main nesting area for the hawksbill turtle (mean 100 ± SD 8.4 clutch-es per year, range 93-111, n = 400 clutches) and the green turtle (mean 9.2 ± SD 6.2, range 1-15, n = 37). Hawksbill turtles nest all through the year, although most nest-ing occurs in the period June-November. The estimated annual number of hawksbill nesting females varied from 21 to 25 turtles. The size of nesting hawksbill turtles in the Dominican Republic (mean CCL = 87.2 cm) is similar to the size of the species in other Caribbean regions.

5) The high level of human predation on clutches and coastal development plans are the main threats for the conservation of these marine turtle nesting stocks. To face egg take, an artificial incubation program was carried out.

6) In the Jaragua National Park artificial in-cubation was carried out in two locations (western and eastern beaches) and hatching success and sex ratio was compared with values of in situ incubated clutches. The results showed that in the west, artificial incubation significantly decreased hatch-ing success in clutches and, in the east, the incubation duration increased, which we predict would result in an increase in male production from these clutches.

7) Clutch relocation is currently the only viable conservation option for clutches on eastern beaches due to intense egg take


11) The sector of La Cueva beach hosts 20% of the total clutches laid on western beaches of the Jaragua National Park and demonstrated the highest hatching success levels. Since this sector is located in the buffer zone out of the Park limits, its inclu-sion inside the limits of the Jaragua Nation-al Park is strongly recommended.

12) Through satellite tracking it was found that during the internesting period hawks-bill turtles remained in the territorial waters of Dominican Republic, mostly over the continental shelf (<200 m) in areas charac-terized by relatively shallow waters close to the corresponding nesting beaches.

13) Overall, the common core-use area of nesting hawksbill turtles of Saona during internesting period was situated inside the Del Este National Park boundaries. Home ranges concentrated in waters at the east-ern-most tip of the island, showing similar location and extension both across turtles and years. Efforts should be increased to mitigate illegal fishing and to restrict boat traffic in these waters.

14) During the foraging period, 78.0 % of locations were outside marine protected areas either in waters of the Dominican Republic and in international waters of Ba-hamas, Nicaragua and Honduras. Our re-sults highlight the significance of different protected areas in the Dominican Republic as hawksbill‘s foraging areas, showing the need of enforcing existing legislation and the expansion of protected areas in the country.

15) The present study also corroborates that the waters off Nicaragua and Honduras are exceptionally important foraging areas for

(~100%), but steps are needed to ensure that natural sex ratio is not distorted. How-ever, on the western beaches, in situ clutch incubation seems possible through beach protection.

8) The artificial incubation program con-ducted in Saona from 2007 to 2010 allowed the release of more than 12000 hawksbill hatchlings. No differences in hatching and emergence success between in situ and ar-tificially incubated clutches were found. However, low temperatures and long in-cubation periods recorded in artificially incubated clutches suggest a bias to male hatchling production. Although artificial incubation is effective in terms of hatchling production, low levels of female production highlights the need to improve this protec-tion measure.

9) We also studied the factors affecting hatching success of leatherback clutches on the western beaches of the Jaragua Na-tional Park to inform possible mitigation through clutch translocation in the case of habitat loss. Beach sector, incubation dura-tion, date of lay and clutch size have been found as the main factors driving hatching success.

10) Clutches of leatherback turtles on the western beaches of the Jaragua National Park presented an unusual high hatching success (75.2 ± 21.0%) for the species as compared to other rookeries in the Carib-bean (~ 50%). Given the exceptional val-ue of hatching success and the current and potential threats affecting leatherback nest-ing beaches, additional efforts in regulation and management of this protected area are needed.

Conclusions I 195

hawksbill turtles nesting in the Caribbean region, showing the turtles vulnerability in these waters.

16) Despite the importance of protected coastal and marine areas, not only for nest-ing marine turtles but also for marine bio-diversity in the country, the management of these areas is limited due to lack of resourc-es and effort. The recovery of nesting colo-nies in the country depends on competent management of protected areas, as well as effective implementation of laws prohibit-ing the consumption and trade of any ma-rine turtle products.

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THIS THESIS IS FOCUSED ON IDENTIFYING THE MAIN MARINE TURTLE NESTING ROOkERIES IN THE DOMINI-CAN REPUBLIC, DESCRIBING THE CURRENT SPATIO-TEMPORAL PATTERNS OF NESTING, AND ASSESSING THE LIkELY IMPACT OF THE CURRENT THREATS TO THESE NESTING STOCkS, SPANNING A 5-YEAR STUDY PERIOD (2006 – 2010).

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